Animal Biodiversity and Conservation issue 35.2 (2012) by Museu Ciències Naturals de Barcelona MCNB

Formerly Miscel·lània Zoològica

2012

and

Animal Biodiversity Conservation 35.2

Dibuix de la coberta: Coturnix coturnix, guatlla, codorniz, quail (Jordi Domènech)

Editor Executiu / Editor Ejecutivo / Executive Editor Joan Carles Senar Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer Assistència Tècnica / Asistencia Técnica / Technical Assistance Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Secretaria de Redacció / Secretaría de Redacción / Editorial Office Museu de Ciències Naturals de Barcelona Passeig Picasso s/n. 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail abc@bcn.cat

Editors / Editores / Editors Pere Abelló Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Javier Alba–Tercedor Univ. de Granada, Granada, Spain Russell Alpizar–Jara Univ. of Évora, Portugal Xavier Bellés Centre d' Investigació i Desenvolupament–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Michael J. Conroy Univ. of Georgia, Athens, USA Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain José Antonio Donazar Estación Biológica de Doñana–CSIC, Sevilla, Spain Jordi Figuerola Estación Biológica de Doñana–CSIC, Sevilla, Spain Gary D. Grossman Univ. of Georgia, Athens, USA Damià Jaume IMEDEA–CSIC, Univ. de les Illes Balears, Spain Jordi Lleonart Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Jorge M. Lobo Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pablo J. López–González Univ. de Sevilla, Sevilla, Spain Juan José Negro Estación Biológica de Doñana–CSIC, Sevilla, Spain Vicente M. Ortuño Univ. de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer IMEDEA–CSIC, Univ. de les Illes Balears, Spain Javier Perez–Barberia The Macaulay Institute, Scotland, United Kingdom Josep Piqué Museu de Ciències Naturals de Barcelona, Barcelona, Spain Oscar Ramírez Inst. de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Montserrat Ramón Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Ignacio Ribera Inst. de Biología Evolutiva CSIC–UPF, Barcelona, Spain Pedro Rincón Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Alfredo Salvador Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Ciències Naturals de Barcelona, Barcelona, Spain Carles Vilà Estación Biológica de Doñana–CSIC, Sevilla, Spain Rafael Zardoya Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Consell Editor / Consejo Editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana–CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle–CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Jersey, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana–CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Barcelona, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway Animal Biodiversity and Conservation 35.2, 2012 © 2012 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: ISSN: 1578–665X Dipòsit legal: B–16.278–58 The journal is freely available online at: www.abc.museucienciesjournals.cat

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XXXth IUGB Congress and Perdix XIII

Editors executius / Editores ejecutivos / Executive Editors Manel Puigcerver Univ. de Barcelona, Spain José Domingo Rodríguez Teijeiro Univ. de Barcelona, Spain Joan Carles Senar Museu de Ciències Naturals de Barcelona, Spain Secretària de redacció / Secretaria de redacción / Managing Editor Montserrat Ferrer Museu de Ciències Naturals de la Ciutadella, Barcelona, Spain

Comité organitzador / Comité organizador / Organizing Comittee Francis Buner Game and Wildlife Conservation Trust, UK Chair of Perdix XIII Manel Puigcerver Univ. de Barcelona, Spain Chair of XXXth IUGB Congress Nicholas Aebischer Game and Wildlife Conservation Trust, UK Antonio Bea Ekos estudios ambientales,Spain. Ricard Casanovas Generalitat de Catalunya, Spain Jorge Cassinello Inst. de Investigación en Recursos Cinegéticos, CSIC, Spain Miguel Delibes Estación Biológica de Doñana, CSIC, Spain Xavier Ferrer Univ. de Barcelona, Spain Francesc Llimona Consorci del Parc de Collserola, Spain José María López Generalitat de Catalunya, Spain Santi Mañosa Univ. de Barcelona, Spain Francesc Piera Federación Catalana de Caza, Spain José Domingo Rodríguez Teijeiro Univ. de Barcelona, Spain Carme Rosell Minuartia, Spain Jordi Ruiz Generalitat de Catalunya, Spain Francesc Sardà Centre Tecnològico Forestal de Catalunya, Spain José Mari Usarraga Federación de Caza de Euskadi, Spain Javier Viñuela Inst. de Investigación en Recursos Cinegéticos, CSIC, Spain

Sessions plenàries / Sesiones plenarias / Plenary Sessions Veterinary aspects of wildlife and conservation: Peter D. Walsh Species extinctions and population dynamics: Philip K. J. McGowan Wildlife law and policy: Borja Heredia Conservation and management of migratory species: Manel Puigcerver Wildlife biology, behaviour and game species management: Nicholas Aebischer Interactions humans–wildlife: Steve Redpath Methodologies, models and techniques: Lisette Waits Human dimensions of game wildlife management: John Linnell

ISSN: 1578–665X

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XXXth IUGB Congress and Perdix XIII F. Buner & M. Puigcerver

Buner, F. & Puigcerver, M., 2012. XXXth IUGB Congress and Perdix XIII. Animal Biodiversity and Conservation, 35.2: 153–154. The 30th Congress of the International Union of Game Biologists (IUGB) and Perdix XIII was held at the Hotel Juan Carlos I in Barcelona, Spain, from 5 to 9 September 2011. The event was organised by the University of Barcelona, the Regional Government of Catalonia Department of Agriculture, Farming, Fish, Food and Environment, the Spanish Institute of Game Resources Research (IREC), and the British Game and Wildlife Conservation Trust. Every two years since the mid–1950s, the International Union of Game Biologists (IUGB) has brought together international wildlife biologists, forestry scientists, veterinarians, game managers, hunters and others with an interest in game or wildlife biology. The IUGB encourages the exchange of scientific and practical knowledge in the field of game and wildlife management, the broad field of game biology, and international co–operation in game and wildlife management. The aim of the conference is to build bridges between scientists, wildlife managers and authorities, and those studying the human dimensions of wildlife management. Following the meetings in Limassol (Cyprus) in 2001 and Braga (Portugal) in 2003, Perdix XIII joined the IUGB Congress series for the third time in its history. Founded in the 1960s, the Perdix series has traditionally attracted partridge, quail and francolin researchers and conservationists from across Europe and North America. To make the Perdix series even more attractive to gamebird biologists, specialists in any Galliform species —whether pheasants, cracids, megapodes or grouse— were welcomed. This joint congress provided a forum to share current developments in gamebird and mammal wildlife research and management, offering an excellent opportunity to identify research gaps, to determine conservation action needs, and to co–ordinate research projects. The congress was attended by 397 researchers and wildlife managers from 37 different countries from the five continents, and included many of the world’s leading wildlife biologists. The general topic was 'Human–wildlife conflicts and peace-building strategies'. The objective was to summarise the general philosophy of the organising and scientific committees to try to overcome the simple collection of problems derived from human–wildlife interactions by proposing solutions on the basis of scientific knowledge of wildlife and management. A total of 260 contributions were presented. Sixty–eight Perdix XIII communications were related to galliform species (38 oral communications and 30 posters). Additionally, keynote plenary lectures were given by renowned experts, each of whom opened one of the eight main topics of the Conference: – –

First plenary session: 'Veterinary aspects of wildlife and conservation' Bushmeat hunting regulates ebola emergence Speaker: Dr. Peter D. Walsh Second plenary session: 'Species extinctions and population dynamics' Galliform species and species extinctions: what we know and what we need to know Speaker: Dr. Philip K. J. McGowan

Francis Buner, Game and Wildlife Conservation Trust, UK. Manel Puigcerver, Dept. de Didàctica de les Ciències Experimentals i la Matemàtica, Fac. de Formació del Professorat, Univ. de Barcelona, Psg. Vall d'Hebron 171, 08035 Barcelona, Espanya (Spain). ISSN: 1578–665X

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– – – – – –

Third plenary session: 'Wildlife law and policy' Policy responses to human-wildlife conflicts. A perspective from the convention of migratory species (CMS) Speaker: Dr. Borja Heredia Fourth plenary session: 'Conservation and management of migratory species' Conservation and management of the Common quail (Coturnix coturnix) in Europe: past, present and future Speaker: Dr. Manel Puigcerver Fifth plenary session: 'Wildlife biology, behaviour and game species management' The Grey partridge in the UK: population status, research, policy and prospects Speaker: Dr. Nicholas Aebischer Sixth plenary session: 'Interactions humans–wildlife' Managing conflicts between conservation and gamebird management Speaker: Dr. Steve Redpath Seventh plenary session: 'Methodologies, models and techniques' Molecular genetic tools and techniques for improving management of wildlife and game species Speaker: Dr. Lisette Waits Eigth plenary session: 'Human dimensions of game wildlife management' Sustainable hunting: an exploration along ecological and social dimensions Speaker: Dr. John Linnell

Of these eight lectures, four were clearly focused on Galliformes species and the others were of general interest to the audience. Six specific workshops were also presented during the Conference, three of which were of particular interest to Perdix attendees: –

Sustainable management of migratory birds – what may hunters and game biologists expect from each other?, led by Dr. Yves Lecocq and Dr. Conor O’Gorman. – GALLIPYR: Pyrenean Network for the mountain game fowl, led by Dr. Virginie Fabre (geieforespir@ forespir.com) and sponsored by the GALLIPYR INTERREG Project. – Reconciling agricultural management, small game production and biodiversity conservation: recommendations for the CAP reform, led by Drs. J. Viñuela, F. Casas, F. Ros, D. Villanúa, P. Ferreras, J. Torres, I. Leranoz, J. Ardaiz, V. Alzaga, A. Cormenzana and E. Castién. Further information can be found on the Conference web page (www.iugb2011.com) where the final programme, the abstract book (in PDF format), and extended abstracts of some contributions can be downloaded. Some of the most outstanding contributions, selected by the scientific committee of the Conference, are now published in this special issue of the international scientific journal Animal Biodiversity and Conservation. We wish to thank the scientific and organising committees, the sponsors, and the participants for making this meeting such an interesting, friendly and highly valuable event.

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XXXth IUGB Congress Comitè científic / Comité científico / Scientific Comittee Wildlife law and policy Sabine Bertouille Lab. Faune sauvage et Cynégétique, Gembloux, Belgium Conservation and management of migratory game species Javier Viñuela Inst. de Investigación en Recursos Cinegéticos, CSIC–UCLM–JCCM, Spain Wildlife biology, behavior and game species management Kjell Sjöberg Swedish Univ. of Agricultural Sciences, Sweden José Domingo Rodríguez Univ. de Barcelona, Spain Interactions humans–wildlife John Bissonette College of Natural Resources, Utah State Univ., USA Carlos M. Fonseca CESAM, Univ. de Aveiro, Portugal Carme Rosell MINUARTIA, Univ. de Barcelona, Spain Methodologies, models and techniques Carles Vilà Estación Biológica de Doñana, CSIC, Spain Antonio Bea Ekos Estudios Ambientales, SLU, Spain Human dimensions of game wildlife management Yves LeCoq Federation of Associations for Hunting and Conservation of the EU, FACE, Brussels Manel Puigcerver Univ. de Barcelona, Spain Assessors dels articles / Asesores de los artículos / Referees of papers Pelayo Acevedo Univ. de Málaga, Spain Beatriz Arroyo Inst. de Investigación en Recursos Cinegéticos, CSIC–UCLM–JCCM, Spain Tyler A. Campbell National Wildlife Research Center, Florida Field Station, USA Carlos Calvete Centro de Investigación y Tecnología Agroalimentaria, Spain Jim Casaer Inst. voor Natuur–en Bosonderzoek, INBO, Belgium Francisca Castro Inst. de Investigación en Recursos Cinegéticos, CSIC–UCLM–JCCM, Spain Isabel Catalán Univ. of Alberta, Canada Adam Dillon Dept. of Fisheries and Wildlife, Virginia Tech., USA Stephen S. Ditchkoff School of Forestry and Wildlife Sciences, Auburn Univ., USA Adriaan Dokter Univ. of Amsterdam, The Netherlands Paulino Fandos Museo Nacional de Ciencias Naturales, CSIC, Spain Carlos M. Fonseca CESAM, Univ. de Aveiro, Portugal Barbara Franzetti Inst. Superiore per la Protezione e la Ricerca Ambientale, ISPRA, Italy Marco Galaverni Inst. Superiore per la Protezione e la Ricerca Ambientale, ISPRA, Italy Germán Garrote Consejería de Medio Ambiente, Junta de Andalucía, Spain José M. Gil Sánchez Spain Alejandro González Estación Biológica de Doñana, CSIC, Spain François Gossman Office National de la Chasse et de la Faune Sauvage, ONCFS, France Jean–Carles Guyomarc’h Univ. Rennes I, France Magnus Johanson Inst. of Ecology and Evolution, Uppsala Univ., Sweden Javier Juste Estación Biológica de Doñana, CSIC, Spain Mike Kaller Louisiana State Univ., USA Oliver Keuling Stiftung Tierärztliche Hochschule Hannover, Deutschland Anne Matilainen Univ. of Helsinki, Findland Maria Miranda Spain Carl Mitchell Wildfowl and Wetlands Trust, UK Alberto Meriggi Univ. di Pavia, Italy Guillem Molina Estación Biológica de Doñana, CSIC, Spain Karen Mustin Univ. of Queensland, Australia Conor O’Gorman The British Association for Shooting and Conservation, UK Joaquín Ortego Inst. de Investigación en Recursos Cinegéticos, CSIC–UCLM–JCCM, Spain Aaron Pearse USGS, Northern Prairie Wildlife Research Centre, USA Eric Petit Univ. Rennes 1/CNRS–UMR 6553 ECOBIO, France Manel Puigcerver Univ. de Barcelona, Spain Lori Randall USGS, Natl. Wetlands Res., USA Dale Rollins Dept. of Wildlife and Fisheries Sciences, San Angelo, USA Pedro Sarmento Portugal David A. Stroud UK Joint Nature Conservation Committee, UK

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XXXth IUGB Índex / Índice / Contents Wildlife law and policy 159–161 S. Bertouille Wildlife law and policy 163–170 K. G. Papaspyropoulos, J. Koufis, L. Tourlida & A. Georgakopoulou Estimating the economic impact of a long term hunting ban on local businesses in a rural areas in Greece a hypothetical scenario

Conservation and management of migratory game species 171–174 M. B. Ellis Management of waterfowl shooting during periods of severe weather in the UK

221–233 S. Cahill, F. Llimona, L. Cabañeros & F. Calomardo Characteristics of wild boar (Sus scrofa) habituation to urban areas in the Collserola Natural Park (Barcelona) and comparison with other locations 235–246 E. Belotti, M. Heurich, J. Kreisinger, P. Šustr & L. Bufka Influence of tourism and traffic on the Eurasian lynx hunting activity and daily movements 247–252 V. J. Colino–Rabanal, J. Bosch, Mª J. Muñoz & S. J. Peris Influence of new irrigated croplands on wild boar (Sus scrofa) road kills in NW Spain

Methodologies, models and techniques

175–188 T. Beroud, J. Druais, Y. Bay & J. C. Ricci Visual counts, bioacoustics and RADAR: three methods to study waterfowl prenuptial migration in Southern France

253–265 K. Morelle, P. Bouché, F. Lehaire, V. Leeman & P. Lejeune Game species monitoring using road–based distance sampling in association with thermal imagers: a covariate analysis

Wildlife biology, behaviour and game species management

267–275 C. Ebert, J. Sandrini, B. Spielberger, B. Thiele & U. Hohmann Non–invasive genetic approaches for estimation of ungulate population size: A study on roe deer (Capreolus capreolus) based on faeces

189–196 C. Fischer & R. Tagand S p at i a l b e h av i o u r a n d s u r v i v a l translocated wild brown hares

197–207 K. Weingarth, C. Heibl, F. Knauer, F. Zimmermann, L. Bufka & M. Heurich First estimation of Eurasian lynx (Lynx lynx) abundance and density using digital cameras and capture–recapture techniques in a German national park 209–217 C. Rosell, F. Navàs & S. Romero Reproduction of wild boar in a cropland and coastal wetland area: implications for management

Interactions humans–wildlife 219–220 C. Rosell & F Llimona Human–wildlife interactions

277–283 M. Narce, R. Meloni, T. Beroud, A. Pléney & J. C. Ricci Landscape ecology and wild rabbit (Oryctolagus cuniculus) habitat modeling in the Mediterranean region 285–293 U. Franke, B. Goll, U. Hohmann & M. Heurich Aerial ungulate surveys with a combination of infrared and high–resolution natural colour images

Human dimensions of game wildlife management 295–306 L. A. Powell Common–interest community agreements on private lands provide opportunity and scale for wildlife management

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Wildlife law and policy S. Bertouille

Bertouille, S., 2012. Wildlife law and policy. Animal Biodiversity and Conservation, 35.2: 159–161. One of the crucial issues of our decades is how to stop the loss of biodiversity. Policy–makers need reliable data to base their decisions on. Managing wildlife populations requires, first of all, science–based knowledge of their abundance, dynamics, ecology, behaviour and dispersal capacities based on reliable qualitative data. The importance of dialogue and communication with the local actors should be stressed (Sennerby Forsse, 2010) as bag statistics and other monitoring data in wildlife management could be more precise if local actors, notably hunters, were better informed and aware of their importance, especially in supporting existing and emerging policies at national and international levels. Another essential issue in wildlife management is the conflicts generated by humans and their activities when they interact with wildlife (Heredia & Bass, 2011). A sociologic approach is required to take into account those human groups whose interests are divergent, facilitating communication and collaborative learning among these users of the same ecosytem. Obstacles should be addressed and solutions devised to protect and encourage a sustainable use of this ecosystem in, as much as possible, a win–win relationship. Policy objectives and management strategies should be discussed and debated among the stakeholders involved, then formulated. Policies can be translated into different types of instruments, economic and legislative, but also informative and educative. As awareness of the actors is a key factor of successful regulation, the regulations should be sufficiently explained and stakeholders should be involved in the implementation of these regulations as much as possible. Finally, the effectiveness of the regulations should be evaluated in light of their objectives, and where necessary, the regulations should be strengthened or adapted to improve their performance (Van Gossum et al., 2010). The various aspects of the processes described above were highlighted in the plenary talk and the five oral communications presented during the session on wildlife law and policy. In his plenary talk, Dr Borja Heredia, Head of the Scientific Unit of the Secretariat of the CMS/UNEP in Bonn, pointed out different sources of human–wildlife conflicts, such as the logging activities in subtropical forests that induce overexploitation and poaching for bushmeat consumption; the problem of predators on livestock and the poisoning of lions in the Masaï Reserve; animals invading the human territory; and game species as a vector of diseases in humans and livestock (Heredia & Bass, 2011). Heredia stressed the importance for wildlife managers to deal with the human dimension; he stressed the importance of successful conflict management based on principles such as a non–adversial framework, an analytical approach, a problem–solving orientation, the direct participation of the conflicting parties, dialogue as a basis for mutual understanding and facilitation by a trained third party. Heredia explained how the Convention on Migratory Species of Wild Animals (UNEP/CMS) contributes to confict resolution and in this way increases the chance of survival of these species. The CMS (see CMS website) works for the conservation of a wide array of endangered migratory animals worldwide through the negotiation and implementation of agreements and action plans. Migratory species threatened with extinction are listed in Appendix I of the Convention. CMS parties strive towards strictly protecting these animals, conserving or restoring the places where they live, mitigating obstacles to migration and controlling other factors that might endanger them. Besides establishing obligations for each State joining the CMS, CMS promotes concerted action among the Range States of many of these species. Migratory species that need, or would significantly benefit from, international co–operation are listed in Appendix II of the Convention. For this reason, the Convention encourages the Range states to reach global or

Sabine Bertouille, Nat. and Agric. Environmental Studies Dept. (DEMNA), Av. Maréchal Juin 23, 5030 Belgium. ISSN: 1578–665X

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regional agreements. The Convention acts, in this respect as a framework convention. The Agreements may range from legally binding treaties (called agreements, there are seven) to less formal instruments, such as Memoranda of Understanding, or actions plans (there are 20), and they can be adapted to the requirements of particular regions. The development of models tailored according to the conservation needs throughout the migratory range is a unique capacity to CMS. Heredia detailed inter alia the Agreement on the Conservation of Albatrosses and Petrels, the Great Apes Survival Partnership, the Agreement on the Conservation of Gorillas and their Habitats, the MoU on the Saïga Antelope, and the Programme for the Conservation and sustainable use of the wild saker falcon (Falco cherrug) in Mongolia. The talk of Sarah Wilks, research fellow at the School of Law, University of Western Sydney, illustrated the importance of adequate transparency and public consultation in environmental and conservation law and decision making. Wilks (2012) examined the Australian legislation concerning animal welfare and the export of Australian wildlife products and, as a case study, explored the Tasmanian State Government’s recent decision to promote the commercial harvest and export of brushtail possums She pointed out that although the Enviromment Protection and Biodiversity Conservation 1999 (EPBC) process intended to be open and co–operative, it is not, in practice, co–operative, public and transparent. The export of possum products requires Australian Government approval under the Department of Primary Industries, Parks, Water and Environment (EPBC). Wilks (2012) assessed the Tasmanian Wildlife Trade Management Plan for Common Brushtail Possums developed by the EPBC, the public submissions to the Australian Government, and the Australian Government’s response against the provisions of the EPBC. As a result, she deplored that welfare outcomes, like that of back or pouch juveniles whose mother had been trapped or killed have not been adequately considered either at Tasmanian State or at Australian Govenment level. She concluded by deploring that submissions on ethical grounds could not yet be considered by the Australian Government because the decision to harvest or not to harvest is made at State level, and yet the Tasmanian State legislation is deficient in mandating public consultation. Data on hunting and game resources provide quantitative and qualitative information on game species, but moreover, game monitoring has shown to be efficient in identifying threats to biodiversity, such as biodiversity problems in agriculture and forest ecosystems, and also to be an early warning in assessing threats from invasive alien species (Sennerby Forsse, 2010). They are an essential tool for game managers, scientists and policy–makers, and hunters and hunter organisations are key resources in the collection of this information.The ARTEMIS data bank was initiated by the Federation of Asssociations of Hunting and Conservation of the European Union FACE (see ARTEMIS website) to improve information about game in support of existing and emerging European policies. The objective of ARTEMIS is to centralise and analyse, in a coordinated and coherent

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way, the information on hunting bags already collected in many European countries and to complete them with new data following a common methodology. As a second step, the Conference on Game Monitoring held in Uppsala, Sweden, in December 2009, aimed to propose further actions to promote streamlined European game monitoring in support of wildlife and biodiversity policies (Sennerby Forsse, 2010). In this context, Martinez–Jauregui & Herruzo (2011) presented data concerning the Spanish hunting statistics collected from 1972 to 2007. Data related to hunters, hunting grounds and game animals were analysed to determine their strengths and weaknesses, and results showed that official Spanish statistics could be incomplete, disperse, and not always homogeneous over a long period of time. The authors concluded that there is a need in the current process to agree on a common international protocol to collect hunting statistics, and they suggested going beyond hunting data to consider other aspects of the hunting sector and reduce the gap between hunting and other agricultural and forest resources. Jutta Gerner, from the Institute of Forest and Environmental Policy at the University of Freiburg, Germany, investigated the shortcomings of the current regulatory practices with regards to hunting regulations in protected areas in order to improve administrative efficiency. Gerner & Schraml (2011), analysed 800 administrative acts and 26 qualitative interviews based on the regulatory arrangement approach (RAA). The RAA is a policy instrument choice theory which helps regulators find the most appropriate instruments by measuring and evaluating them. Van Gossum et al. (2010) developed the RAA by merging current smart regulation theory with the policy arrangement approach and the policy learning concept. Gerner & Schraml (2011) suggested the integration of a more cooperative, less 'regulator' and more informative policy style in hunting regulations. They recommended better communication and information among the concerned administrative sections and between administration and local actors in order to improve policy success. They recommended that the different stakeholders should be informed and involved when debating policy objectives and strategies, as well as in the application of the administrative acts. The study presented by R. Mateo from the 'Instituto de Investigacion en Recursos Cinegeticos', IREC (CSIC, UCLM, JCCM), Spain, was an example of a rigorous follow–up of a regulation objective, followed by a reinforcement and adaptation of the regulation to improve its performance. Results of Mateo et al. (2011) showed that, although the use of lead shot was banned in protected wetlands in Spain in 2001, ban compliance was insufficient, as in 2007–2008 a large number of waterfowl hunted in wetland still had embedded lead shot. After these results were produced, the ban was reinforced and compliance subsequently increased. Nevertheless, in 2009–2010, the last year of this study, a significant proportion of birds still had embedded lead shot and /or ingested lead shot in their gizzards. The authors suggested this occurred because the majority of ducks often feed in unprotected rice fields. They therefore recommended

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extending the ban to all waterfowl hunting and not only that undertaken in protected wetlands. The presentation of K. E. Skordas, from the Hunting Federation of Macedonia and Thrace, Research Division, Greece, illustrated the contribution of the Hellenic Hunters Confederation (HHC) to law enforcement for wildlife protection. It showed how stakeholders, hunters, set up heir own Game Warden Service in 1999, through their Hunting Associations, in order to assume responsibility for the control of illegal hunting and wildlife protection, in collaboration with the local Forest Service. These game wardens carry out repressive and preventive controls and prosecutions. Besides this initiative, information campaigns are organised by the HHC to improve hunters’ awareness (see website of the Hellenic Hunters Confederation, HHC). Skordas & Papaspyropoulos (2011) analysed the relation between law enforcement, hunter awareness and infringement categories, classed in degree of influencing wildlife protection. They observed a strong reduction in the number of infringements; particularly, they found that hunting out of season and hunting without a license decreased from 23.4% to 7.31% and from 30.12% to 11.8%, respectively. All the talks presented in this session stressed the importance of dialogue in wildlife management as a basis for mutual understanding. Communication and involvement of the local actors/stakeholders are key factors at different stages of wildlife management: when collecting reliable data on which policy–makers may draw up their decisions, when debating policy objectives and strategies, and when implementing regulations and administrative acts. References ARTEMIS website available at http://www.artemis– face.eu Accessed 22 March 2012. CMS website, available at http://www.cms.int/about/ index.htm. Accessed 22 March 2012. Gerner, J. & Schraml, U., 2011. From command and

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control to cooperation. Towards an adequate policy mix in hunting regulations in protected areas. Abstract book of the XXXth IUGB Congress and Perdix XIII, 5–9 September 2011, Barcelona: 181. Hellenic Hunters Confederation website available at http://www.ksellas.gr/index_en.asp Accessed 22 March 2012. Heredia, B. & Bass, J., 2011. Policy responses to human–wildlife conflicts. A perspective from the Convention on Migratory Species (CMS). Abstract book of the XXXth IUGB Congress and Perdix XIII, 5–9 September 2011, Barcelona: 32. Martinez–Jauregui, M. & Herruzo, A. C., 2011. A contribution to the current debate on the improvement of hunting statistics: the case of Spain. Abstract book of the XXXth IUGB Congress and Perdix XIII, 5–9 September 2011, Barcelona: 183. Mateo, R., López–Antia, A., Taggart, M.A., Martinez–Haro, M. & Guitar, R., 2011. Lead shot ban compliance in Spanish wetlands: effects on Pb poisoning prevalence. Abstract book of the XXXth IUGB Congress and Perdix XIII, 5–9 September 2011, Barcelona: 184. Sennerby Forsse, L., 2010. Providing a knowledge basis for sustainable hunting and biodiversity conservation. Steamlining hunting bag statistics in the EU. Informal report of the workshop held during the EU–Conference on Game Monitoring, 15–16 December 2009, Upssala, Sweden. Skordas, K. E. & Papaspyropoulos, K. G., 2011. Contribution of Greek Hunting Associations to law enforcement for the wildlife protection. Abstract book of the of the XXXth IUGB Congress and Perdix XIII, 5–9 September 2011, Barcelona: 186. Van Gossum, P., Arts, B. & Verheyen, K., 2010. From 'smart regulation' to 'regulatory arrangements'. Policy Sciences, 43: 245–261. Wilks, S., 2012. Animal welfare and the export of Australian wildlife products: how well does the legislation function. Abstract book of the XXXth IUGB Congress and Perdix XIII, 5–9 September 2011, Barcelona: 70.

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Estimating the economic impact of a long–term hunting ban on local businesses in rural areas in Greece: a hypothetical scenario K. G. Papaspyropoulos, J. Koufis, L. Tourlida & A. Georgakopoulou

Papaspyropoulos, K. G., Koufis, J., Tourlida, L. & Georgakopoulou, A., 2012. Estimating the economic impact of a long–term hunting ban on local businesses in rural areas in Greece: a hypothetical scenario. Animal Biodiversity and Conservation, 35.2: 163–170. Abstract Estimating the economic impact of a long–term hunting ban on local businesses in rural areas in Greece: a hypothetical scenario.— In December 2009, hunting was banned for a few days in Greece following the decision of the Council of State. The decision was issued when an animal rights organization claimed to the Court that there was no updated evidence about the impact of hunting on wild populations. This case prompted the present study, which focused on examining the hypothetical scenario of the possible impact of a long–term hunting ban on local businesses in rural areas in Greece. We carried out face–to–face interviews with entrepreneurs from the accommodation and food service sectors. Our results showed that most business owners interviewed considered the impact would be significant for their annual earnings. This finding should be taken into account by environmental decision makers because rural and mountainous areas in Greece are sparsely populated, and the few small businesses that still operate would not withstand drastic changes in rural tourism. Key words: Economic contribution of hunting, Hunting restrictions, Accommodation services sector, Food service sector. Resumen Estimación del impacto económico de una veda de caza a largo plazo sobre los negocios locales en las zonas rurales de Grecia: una situación hipotética.— En diciembre del 2009, en Grecia se prohibió la caza durante unos pocos días, siguiendo la decisión del Consejo de Estado. Esta se tomó cuando una organización defensora de los derechos de los animales recurrió a la Corte argumentando que no existían pruebas actualizadas sobre el impacto de la caza sobre las poblaciones de animales salvajes. Estas circunstancias promovieron el presente estudio, que se enfocó hacia el examen de unas hipotéticas circunstancias del posible impacto de la veda de caza a largo plazo sobre los negocios locales de las zonas rurales de Grecia. Llevamos a cabo entrevistas cara a cara con los empresarios de los servicios de alojamiento y gastronomía. Nuestros resultados mostraron que la mayoría de propietarios de negocios entrevistados consideraban que el impacto sería significativo para sus ingresos anuales. Los gestores del medio ambiente deberían tener en cuenta este resultado, dado que las áreas rurales montañosas de Grecia están escasamente pobladas, y los pocos negocios que aún funcionan en ellas no podrían soportar cambios drásticos en el turismo rural. Palabras clave: Contribución económica de la caza, Veda, Sector de servicios hoteleros, Sector de servicios gastronómicos. Received: 9 III 12; Conditional acceptance: 2 IX 12; Final acceptance: 1 X 12 Konstantinos G. Papaspyropoulos, Lab. of Forest Economics, Fac. of Forestry and Natural Environment, Aristotle Univ. of Thessaloniki, Univ. Campus, P. O. Box 242, GR–54124 and Hunting Federation of Macedonia and Thrace, Research Division, Ethnikis Antistaseos 173–175, 55134, Kalamaria, Thessaloniki, Greece.– John Koufis, Lamprini Tourlida & Anastasia Georgakopoulou, Dept. of Forestry and Natural Environment Management, Technological Education Inst. of Lamia, Karpenissi Anexx, 36100 Karpenissi, Greece. Corresponding author: K. G. Papaspyropoulos. E–mail: kodafype@for.auth.gr ISSN: 1578–665X

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Introduction Hunting activity contributes to the economies of rural and mountainous areas (Booth, 2010; Samuelsson & Stage, 2007; Papaspyropoulos et al., 2012) and relatively high revenues can be generated by few clients (CIC, 2008; Lindsey et al., 2007). In Greece, the hunting period lasts from August 20th to March 10th (Hellenic Ministry of Environment, Energy, and Climate Change, 2012). During this period, business sectors close to hunting sites, such as those in the accommodation and food service sectors, are influenced financially by the hunting activity. The accommodation sector generates more than 1.3 million €, while the food service sector generates almost 7.3 million €, according to an independent survey ordered by the Panhellenic Union of Hunting Material Merchants (PEVEKE, 2011). At the same time, there is a strong anti–hunting movement in Greece. Many animal rights organizations, environmental NGOs, and other groups undertake actions trying to restrict or completely ban hunting activity. During one such event, in 2009, the 'Greek Animal Rights and Ecological Association' was heard by the Hellenic Council of State for judicial review of the annual act which regulates hunting in Greece. The Council of State agreed the act needed updating and it ordered a hunting ban until this was undertaken. In reality, the ban lasted only a few days because an updated study about the abundance of game populations in Greece was submitted to the Ministry of Environment. The idea of the present research paper was prompted by this incident, and by the fact that the anti–hunting movement is quite strong in Greece, and there are constant conflicts between hunting organizations and groups trying to restrict or ban hunting activity. We therefore set up a hypothetical scenario to examine how entrepreneurs considered their business would be affected if at some time a long–term hunting ban was imposed on a local area, or generally, in Greece. In this study we assessed the economic impact of a hypothetical hunting ban on small local businesses in terms of mean percentage of the total annual income of businesses, based on the entrepreneurs’ testimonies. Background Economic contribution of hunting Hunting plays a significant role in the economy of many countries (Molina–Martinez et al., 2002; Sokos et al., 2003; Croitoru, 2007; Sokos et al., 2009). In most northern Mediterranean countries, hunting benefits range from €4–8/ha, while for most southern and eastern Mediterranean countries estimated values are within €1–4/ha (Croitoru, 2007). In Greece, hunting contributes 2.3 billion €/annum to the country’s economy (PEVEKE, 2011). A significant part of this amount can be allocated to hunting tourism (accommodation and food service sectors). Hunters spend a mean of 558€/ annually and 58 €/hunting trip per person for accommodation and dining, respectively.

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Many hunters, however, prefer to travel abroad to hunt, contributing significantly to the economies of the host countries. For example, the gross value of hunting tourism in South Africa was estimated at 68.4 million $ in 2003 (Booth, 2010), and 30 million $ in Canada (MacKay & Campbell, 2004). Overall, in Europe and the USA hunting revenues are estimated at 16 billion €/annum and 76 billion $/annum, respectively (Booth, 2010; Grado et al., 2011). Hunting and anti–hunting in Greece There are approximately 220 thousand hunters in Greece. Hunting is regulated by the Hellenic Ministry of Environment, Energy and Climate Change (2012). According to the Forest Law of 1969, the quarry belongs to the hunter and not to the land owner, a quite unique situation compared to other countries. A hunting license is required and is valid for a prefecture, a geographical region, or the whole country. The hunting licence fee is 100–150€. This money is then allocated to the Hunting Organizations and the public 'Green Fund' for the management of hunting activity. It is the hunting organizations that mainly employ wildlife ecologists and wardens and finance hunting management. The contribution of the Green Fund through Forest Service to hunting development is low (Birtsas et al., 2009). Additionally, hunting organizations, as institutional actors in hunting, put pressure on the Government for fundamental hunting issues such as hunting restrictions. Pressure on the Government is also exerted by the the anti–hunting movement in Greece, which is guided by animal rights organizations, environmental NGOs, political movements and other actors. This movement seems to present a strong urban character, similar to movements in other countries (Duda & Jones, 2009). The anti–hunting movement peaked in December 2009 when the Hellenic Council of State decided that the petition of the 'Greek Animal Rights and Ecological Association' against the annual hunting law was fair and that there was not available updated evidence about the abundance of wild populations of game species in Greece. The following analysis was prompted by the fact that the conflicts between hunting organizations and the anti–hunting movement may sometime in the future result in a local or general long–term hunting ban. Methodology Data were collected through structured, face–to–face interviews conducted in the autumn of 2010. According to Maughan et al. (2004), 'the advantages of this approach for quantitative studies are that researchers can feel confident that the same ‘stimulus’ has been presented to all study participants, interviewer effects are minimised and, provided the questions are well– worded, good reliability should be relatively easy to achieve'. The questionnaires were administered to businesses in three administrative regions (former prefectures) Evritania, Messinia and Aitoloakarnania

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B Greece Salonika

Messologi

Athens

Kalamata

Karpenisi

Fig. 1. Research area: A. Evritania Prefecture; B. Aitoloakarnania Prefecture; C. Messinia Prefecture. Fig. 1. Área de investigación: A. Prefectura Evritania; B. Prefectura Aitoloakarnania; C. Prefectura Messinia.

(fig. 1). Evritania is a winter destination for domestic tourists (accommodation capacity peaks in December and January), while Messinia and Aitoloakarnania are mostly summer destinations for domestic and outbound tourists (accommodation capacity peaks in July–August) (Hellenic Statistical Authority, 2012) Two types of questionnaires were administered; one for the accommodation service sector and one for the food service sector. The businesses were selected from the business catalogs obtained from the 'Chambers of Commerce and Industry' in each of the three regions (six catalogs). The catalogs were merged into two (one for each sector). A simple random sample was then taken from every catalog with the use of a random number table (Fowler & Cohen, 1995), and 74 businesses from the accommodation sector (20% of the population) and 89 from the food service sector (5% of the population) participated in the research. The uneven percentages of samples are due to the fact that the businesses in the food

service sector were far more numerous than those in the accommodation sector. Because all the interviews were personal (all three researchers visited the businesses, and they had to revisit some of them in order to interview the entrepreneurs), all the participants responded; therefore, the response rate was 100%. Table 1 shows the variables used in the survey; these were used as the main questions during the interviews. The statistical analysis was performed using basic descriptive statistics, such as frequencies and means (Bradley, 2007) and the multivariate statistical method Correspondence Analysis. According to Markos et al. (2010) 'Correspondence Analysis is a multidimensional data analytic method, suitable for graphically exploring the association between two or more, non–metric variables without a priori hypotheses or assumptions'. The theoretical foundations of this method can be found in Papadimitriou (2007) and Greenacre (2010). The nominal variable 'Administrative region' and the ordinal one 'Do you consider this turnover significant for your

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Table 1. Variables used in the survey. Tabla 1. Variables utilizadas en el estudio.

Variable

Type

Levels

Administrative region (prefecture)

Nominal

Evritania

Messinia

Aitoloakarnania

Type of business

Accomodation sector

Nominal

Food service sector

Do hunters use the business?

Yes

Nominal

Price of the service

Scale

–

Perception about hunting

Nominal

Agree with hunting

Neutral

Disagree with hunting

Perception about hunter as client

I want the hunter as a client

Nominal

Neutral

I do not want hunter as a client

Turnover percentage by hunters as clients

Scale

Do you consider this turnover significant

Ordinal

Important

for your business?

Neutral

Negligible

Would a hunting ban result in an economic

Yes

Nominal

impact on your business?

business?' were used in the Correspondence Analysis. The reason for using such a technique was that it can visualize data of a contingency table of two categorical variables and reveal patterns not apparent on the table frequencies (Greenacre, 2010). The variables that were used were determinant for revealing if the perceived category of economic impact (important, neutral, negligible) is related to a specific Greek rural area. Prior to performing correspondence analysis, the two variables were tested for collinearity with the Variance Information Factor, after they had been transformed to dummy variables (Hair et al., 2006). All statistical analyses and the model validation were performed and confirmed using SPSS 19.0 (Kinnear & Gray, 2011). Results Some descriptive statistics from the business owners’ answers were extracted first. Table 2 shows the absolute value and the percentage of businesses in the accommodation and food service sectors used by hunters in the three administrative regions. It also shows the mean price per service offered by the two sectors in the corresponding regions.

It was found that most businesses included in the sample are used by hunters. Businesses in the Evritania region, especially those in the accommodation sector, seem to be used less. This can be explained by the fact that Evritania is a region that is more than four hours drive away from Athens and Thessaloniki (the two largest cities in Greece and those with most hunters). Therefore, hunters may not choose it very often as a hunting place. Furthermore, Evritania is a mountainous area which, during the hunting period, is also influenced by other forms of tourism, such as adventure sports tourism, or religious tourism. Accommodation is expensive, and maybe hunters choose not to stay overnight in the region. Table 3 shows owners´ perceptions about hunting and about hunters as clients. It shows a similar pattern to table 2. The Evritania region, especially concerning its accommodation sector, seems to differ from the other two regions. Less than half the businesses in Evritania’s accommodation sector seem to support hunting activity (slightly higher in the food service sector). However, the relation between the two variables was significant (Phi & Cramer’s V p–value < 0.05), which means that those who wanted hunters as clients were also in agreement with hunting activity. Only 4.3% of the entrepreneurs were against hunting: however all of them wanted hunters as clients.

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Table 2. Hunters’ use of local businesses and businesses’ mean price per service: Uh. Used by hunters; Mr. Mean price per room; Mp. Mean price per person. Tabla 2. Uso por parte de los cazadores de los negocios locales y precio medio por cada servicio: Uh. Utilizado por los cazadores; Mr. Precio medio por habitación; Mp. Precio medio por persona.

Accommodation sector

Food service

Total

Evritania

12 (63%)

68.5

26 (87%)

15.0

Messinia

28 (93%)

35.0

28 (97%)

12.6

Aitoloakarnania

19 (76%)

38.0

30 (100%)

12.2

Table 3. Enterpreneurs’ perceptions about hunting and hunters as clients: Ah. Agree with hunting; Hc. I want hunter as a client. Tabla 3. Percepciones de los propietarios de los negocios sobre la caza y los cazadores como clientes: Ah. Estoy de acuerdo con la caza; Hc. Quiero al cazador como cliente. Accommodation sector

Food services sector

Total

Evritania

8 (42%)

14 (74%)

16 (53%)

23 (77%)

Messinia

22 (73%)

25 (83%)

22 (76%)

25 (86%)

Aitoloakarnania

15 (60%)

25 (100%)

22 (73%) 30 (100%)

Table 4. Economic impact of hunting ban and its significance for the entrepreneurs: Ma. Mean annual income (in %). Ts. Do you consider this turnover significant for your business? (answering yes); Hb. Would a hunting ban result in an economic impact on your business? (answering yes). Tabla 4. Impacto económico de la veda de caza y su importancia para los propietarios de los negocios: Ma. Ingreso anual medio (en %); Ts. ¿Considera este intercambio significativo para su negocio? (si como respuesta; Hb. ¿La prohibición de caza afectará economicamente a su negocio? (sí como respuesta).

Accommodation sector

Food services sector

Evritania

6.4

21%

63%

6.3

33%

73%

Messinia

8.6

39%

90%

11.5

14%

90%

Aitoloakarnania

20.0

84%

100%

17.5

95%

100%

Table 4 shows the answers to the main question in this study: how much and how significantly would a hunting ban affect local businesses? The table presents the estimates from the owners' mean annual income from hunters, its significance compared to their total annual income, and, if eliminated due to a hunting ban, whether this would have a significant economic impact on their businesses. It shows that the mean annual in-

come that would be lost for the businesses in the three regions varies from 6.3 to 20.0%. All three regions and both sectors believe that this would be an economic impact for their operation. However, especially in the Evritania region, it seems that there are few businesses which consider this turnover significant. This seems to confirm the previous finding, which showed that in Evritania, where the two sectors can rely on other

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Table 5. Results of Correspondence Analysis. Tabla 5. Resultados del Análisis de Correspondencias. 'Administrative unit' x 'Significance of economic impact' Dimension

Inertia

Variance explained

Chi–square

p–value

First

0.36

85.8

252 (df = 4)

0.000

Second

0.06

14.0

forms of winter tourism, the impact of a hunting ban is considered important but not as significant as the turnover from other activities. The above findings also seem to be confirmed by the application of the Correspondence Analysis. This methodology revealed that most of the information (variance) of the model is explained at the first dimension, thus the one–dimension solution is the best one. More than 85% of the total variance is explained, giving a good picture of the relation between the variables of 'Administrative region' and 'Significance of economic impact'. A little variation is explained by the second dimension, not more than 14%. Table 5 shows the inertia value of the first two dimensions, the variance explained, and the chi–square statistic, which justifies the assumption that the two variables are related. However, there was no collinearity because the Variance Inflation Factor had a value < 2 in all cases.

Figure 2 shows the interaction of the two variables. This figure reveals a pattern for the relation between the administrative region and the perception of the business owners in these regions about the economic impact of a hunting ban. It indicates that Aitoloakarnania’s accommodation and food service sectors consider the economic loss as important for their operation. Messinia and Evritania, if seen in the first dimension, consider this impact as neutral or negligible. However, in the second dimension, it is only Evritania which is seen to consider the impact as negligible. Discussion The research confirmed that local businesses are used by hunters during their hunting trips. Entrepreneurs understand the economic contribution of the activity

1.5

Prefecture Significance of economic impact

Dimension 2

1.0 0.5

Negligible Evritania

Important

0.0

Aitoloakarnania

–0.5

Messinia Neutral

–1.0

–1.5 –1.5

–1.0

–0.5 0.0 0.5 Dimension 1

Fig. 2. Interaction of the two variables. Fig. 2. Interacción de las dos variables.

1.0

1.5

Animal Biodiversity and Conservation 35.2 (2012)

and rely on income from hunting. They see hunting as positive, and they understand that a long–term ban may worsen their business’ financial position. Our results imply that in a region where winter tourism is not especially popular, such as the Aitoloakarnania region, the economic impact of a hypothetical long–term hunting ban is considered important. Hunters support the viability of small local businesses through their activity, and they cannot rely on other forms of tourism in the winter. On the other hand, in a region where the income in the accommodation or the food service sectors relies on tourism in a particular season, such as ski–tourism, then hunting tourism, and hunting in general seems necessary but not determinant for the viability of the local businesses. Evritania and Messinia are two such regions; the former in the winter, and a little bit in the summer, and the latter especially in the summer. This pattern confirms previous studies in Greece which report that hunting supports small businesses in rural areas, especially in winter, where there are no other visitors and few other potential revenue sources (Sokos et al., 2003; Tsachalides et al., 2003; Hasanagas et al., 2008). Of course, the present results cannot be generalized for the whole of Greece. Other regions would need to be included in the sample before more general conclusions can be extracted. It could be expected, however, that the pattern would be confirmed if rural regions in northern Greece were included in the research as such areas are not characterized as attractive to commercial tourism, especially if compared to the Greek islands or other renowned regions. Future studies could be conducted to estimate such an impact in terms of jobs lost, in terms of businesses closing down, or in terms of people leaving their homeland to find new jobs. The impact on other sectors —such as the energy sector and the hunting merchandise sector— could also be estimated. Findings from such studies could translate all these effects into actual amounts of money, and not percentage estimations, which is a limitation of the present paper. Results from the present study suggest that a long–term hunting ban would have a significant economic impact (6.3–20%) on businesses that depend on hunting activity. It should be kept in mind that there are instances when hunting bans did not achieve the main objectives that they were set out to meet, that is, to increase populations and eliminate poaching. Baker et al. (2002), for example, found that the hunting ban of foxes in Britain had no measurable impact on fox populations. Additionally, in a study of sub–Saharan Africa, Lindsey et al. (2007) found that as well as a loss in revenue, a hunting ban led to an upsurge in poaching due to the removal of incentives for conservation. This latter finding is also apparent in Greece, in the Amvrakikos wetland. Athough a hunting ban has been in place for many years in this region, both environmental NGOs (which are still in favour of the ban) and hunting organizations (which are against the ban) complain about the poaching in the region, illustrating how a hunting ban can impact on species abundance policy as well as on the local economy. Therefore, the state should develop and implement more integrated management plans for wildlife, also

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taking into account the local people needs and the socioeconomic impacts of hunting bans on small local businesses in rural areas. Acknowledgements We would like to thank the Editor and the anonymous reviewers for their constructive comments. References Baker, P. J., Harris, S. & Webbon, C. C., 2002. Ecology: Effect of British hunting ban on fox numbers. Nature, 419: 34. Birtsas, P. K., Sokos, C., Hasanagas, N. & Billinis, C., 2009. The hunting activity in Hellas. In: Proc. VIth International Symposium on Wild Fauna. Organized by Wild Animal Vigilance Euromediterranean Society 21–24/5/2009 Paris. Extended abstracts: 52–53. Booth, V. R., 2010. The Contribution of Hunting Tourism: How Significant is This to National Economies? In: Contribution of Wildlife to National Economies. Joint publication of FAO and CIC. Budapest. Bradley, T., 2007. Essential statistics for economics, business and management. John Wiley & Sons, Chichester, England. CIC, 2008. Best Practices in Sustainable Hunting. A Guide to Best Practices From Around the World. CIC Technical Series Publication No. 1, Ministry of Environment, 2011. Croitoru, L., 2007. How much are Mediterranean forests worth, Forest Policy and Economics, 9: 536–545. Duda, M. D. & Jones, M., 2009. Public Opinion on and Attitudes toward Hunting. Transactions of the Seventy–third North American Wildlife and Natural Resources Conference. Wildlife Management Institute, Washington D.C. Fowler, J. & Cohen, L., 1995. Statistics for Ornithologists. Edition 2. BTO Guide 22. Grado, S. C., Hunt, K. M., Hutt, C. P., Santos, X. T. & Kaminski, R. M., 2011. Economic Impacts of Waterfowl Hunting in Mississippi Derived From a State–Based Mail Survey, Human Dimensions of Wildlife, 16.2: 100–113. Greenacre, M., 2010. Biplots in practice. Fundacion BBVA, Barcelona. Hasanagas, N., Birtsas, P. & Sokos, C., 2008. Social characteristics and aesthetic ecotourism of hunters. 3rd Environmental Conference of Macedonia. Greek Chemists Union. Thessaloniki 14–17/3/2008. http://www.panida.gr/area/wp-content/uploads/2008_ social-characteristics-and-ecotourism-of-hunters.pdf Hair, J. F., Black., W. C., Babin, B. J., Anderson, R. E. & Tattum, R. C., 2006. Multivariate Data Analysis. Sixth edition. Pearson Prentice Hall, Upper Saddle River, N.J. Hellenic Ministry of Environment, Energy, and Climate Change, 2012. Hunting regulation for the hunting period 2012–2013. Athens, Greece. [In Greek.] Hellenic Statistical Authority, 2012. http://www.statis-

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tics.gr (accessed on 16 September 2012). Kinnear, P. R. & Gray, C. D., 2011. IBM SPSS 19 Made Simple. Psychology Press, East Sussex, UK. Lindsey, P. A., Roulet, P. A. & Romañach, S. S., 2007. Economic and conservation significance of the trophy hunting industry in sub–Saharan Africa. Biological Conservation, 134(4): 455–469. MacKay, K. J. & Campbell, J. M., 2004. An examination of residents’ support for hunting as a tourism product. Tourism Management, 25(4): 443–452. Markos, A., Menexes, G. & Papadimitriou, I., 2010. The CHIC Analysis Software v1.0. In: Classification as a Tool for Research: 409–416 (H. Loracek– Junge & C. Weihs, Eds.). Proceedings of the 11th IFCS Conference, Springer, Berlin. Maughan, B., 2004. Investigator–based interviews. International Journal of Market Research, 46(1): 99–102+122. Molina–Martinez, J., Vinuela, J. & Villafuerte, R., 2002. Reconciling gamebird hunting and biodiversity (REGHAB). Socio–economic aspects of gamebird hunting, hunting bags, and assessment of the status of gamebird populations in REGHAB countries. Part 1: Socioeconomic and cultural aspects of gamebird hunting. REGHAB, London. Papadimitriou, I., 2007. Data Analysis: Correspondence Analysis, Hierarchical Cluster Analysis and other methods. Typothito publications, Georgios Dardanos, Athens. [In Greek.] Papaspyropoulos, K. G., Sokos C. K., Hasanagas N.

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D. & Birtsas, P. K., 2012. Sustainability of recreational hunting tourism: a cluster analysis approach for woodcock hunting in Greece. In: New Trends Towards Mediterranean Tourism Sustainability. Nova Science Publishers. PEVEKE, 2011. Hunting and its contribution to Greek society and economy today. MRB. Samuelsson, E. & Stage, J., 2007. The size and distribution of the economic impacts of Namibian hunting tourism. South African Journal of Wildlife Research, 37(1): 41–52. Sokos, C., Hasanagas, N., Papaspyropoulos, K. G. & Birtsas, P., 2009 Hunting management and hunting – related values. Proceedings of 2nd Conference on Environmental Management, Engineering, Planning and Economics (CEMEPE 2009). http://www.panida.gr/site/wp-content/uploads/2009hunting-management-values.pdf. Sokos, C. Κ., Skordas, Κ. Ε. & Birtsas, P. Κ., 2003. Valuation of hunting and management of brown hare (Lepus europaeus) in rangelands. Proceedings of 3rd Pan–Hellenic Rangelands Conference. Hellenic Rangelands and Pasture Society, 4–6 September 2002, Karpenissi, Greece. (In Greek.) http://www.panida.gr/site/wp-content/uploads/2003harehunters.pdf. Tsachalidis, E., Galatsidas S. & Tsantopoulos, G. E., 2003. Personal characteristics of the hunters in the hunting association at Macedonia and Thrace, North Greece. Ann. Scient. Forest. Dept. Vol. N. [In Greek.]

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Management of waterfowl shooting during periods of severe weather in the UK M. B. Ellis

Ellis, M. B., 2012. Management of waterfowl shooting during periods of severe weather in the UK. Animal Biodiversity and Conservation, 35.2: 171–174. Abstract Management of waterfowl shooting during periods of severe weather in the UK.— During prolonged periods of severe weather waterfowl habitats are affected by snow or ice. Recreational users of the countryside can move waterbirds off prime feeding areas, potentially resulting in reductions in body condition, at a time when energy reserves are key to overwinter survival and subsequent breeding success. Over the last 30 years the British Association for Shooting and Conservation has been closely involved, along with the government and other conservation NGOs, in developing and implementing a criteria–driven process for defining severe weather and managing waterfowl shooting during it in order to minimise unnecessary disturbance to overwintering waterbirds. The system works well and is widely respected and applauded by conservation agencies. There are increasing efforts to bring more countryside activities, including dog walking, bird watching/ringing and other wetland users, into the system so as to reduce further unnecessary waterfowl disturbance during difficult winter weather. Key words: Waterfowl, Management, Hunters, Weather. Resumen Gestión de la caza de aves acuáticas durante los periodos de condiciones meteorológicas adversas en el Reino Unido.— Periodos prolongados de condiciones meteorológicas muy adversas provocan que los hábitats de las aves acuáticas se vean afectados por la nieve o el hielo. El uso recreativo de estas zonas puede desplazar las aves fuera de sus áreas de alimentación principales, lo que potencialmente ocasionaría una reducción de sus condiciones físicas en un momento en que las reservas energéticas son claves para la supervivencia invernal y el subsecuente éxito reproductivo. Durante los últimos 30 años la Asociación Británica para la Caza y la Conservación se ha implicado mucho en este problema, junto con el gobierno y otras ONGs ecologistas, desarrollando y aplicando un criterio que defina, en función de los datos, condiciones meteorológicas muy adversas y la gestión de la caza de aves acuáticas durante este periodo, para minimizar las molestias innecesarias a dichas aves, que están invernando. Este sistema funciona bien y es muy respetado y aplaudido por las instituciones conservacionistas. Se están haciendo cada vez más esfuerzos para establecer más actividades en el espacio natural, incluyendo el paseo de perros, la observación/anillado de aves y otros usos de los humedales, con el fin de reducir las perturbaciones innecesarias a las aves acuáticas durante el difícil tiempo invernal. Palabras clave: Aves acuáticas, Gestión, Cazadores, Tiempo meteorológico. Received: 19 XII 11; Conditional acceptance: 10 II 12; Final acceptance: 23 III 12 Matthew B. Ellis, British Association for Shooting and Conservation, Marford Mill, Wrexham Road, Rossett, Wrexham, LL12 0HL, Wales, UK.

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Introduction During winter waterbirds can be placed under extra stress due to the joint demands of increasing thermoregulation and foraging in areas of diminishing food reserves (Goss–Custard et al., 1977; Clark, 2002). In severe winters this can lead to the depletion of fat reserves and eventually to death (Camphuysen et al., 1996; Suter & Vaneerden, 1992; Davidson & Evans, 1982). Disturbance of these birds, either by predators (Cresswell & Whitfield, 2008; Whitfield, 2003; Hilton et al., 1999) or human activity (Gill et al., 2001; Tamisier et al., 2003; Burton et al., 2002; Burger, 1998) can act to exclude birds from their feeding areas, as well as causing them to expend energy in their escape flights, potentially leading to greater risks of mortality in these populations. The open season for waterfowl shooting in the UK starts on 1 September, and finishes on 31 January at inland sites, but continues until 20 February in coastal areas between the low and high watermarks (the foreshore) (Wildlife and Countryside Act 1981 s2(4)). The Wildlife and Countryside Act 1981 makes provisions for the Secretary of State in the relevant country to make a Protection Order (Wildlife and Countryside Act 1981 s2(6)), in effect imposing a temporary closed season, primarily in order to protect birds during periods of prolonged severe weather. The following is a brief description of the system, its legal basis and the important role that non–governmental organisations, including hunting and conservation bodies, have in managing the process and informing their members and the wider public. This system has been in place since 1967 (The Protection of Birds Act, 1967), but was first used in 1979 (for a full history of the management of waterfowl shooting during periods of severe weather see Stroud et al. (2006) and Stroud (1992). As the largest hunting organisation in the UK the British Association for Shooting and Conservation (BASC) has been involved with the process from the start, along with other NGOs including the Game and Wildlife Conservation trust (GWCT), the Royal Society for the Protection of Birds (RSPB), the Wildfowl and Wetlands Trust (WWT) and the British Trust for Ornithology (BTO). The system has been reviewed on numerous occasions, and continues to be reviewed after any winter in which the severe weather process is used. All of the countries of the United Kingdom operate a severe weather process. However, England and Wales operate as a single severe weather region, but work very closely with Scotland, which is also a separate severe weather region. The Northern Ireland system operates separately from the rest of the UK, but in recent years has developed a very similar process. Severe weather criteria Each day from 9 November to 20 February the UK Meteorological Office (Met office) collects the minimum air and grass temperatures from the previous twenty–four hours across the network of meteorological

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stations (fig. 1) and reports the data to the Joint Nature Conservation Committee (JNCC). These stations are mostly coastal to reflect the fact that inland waters in the UK tend to freeze first, forcing birds to the coast. In Northern Ireland the Met office deals directly with the Northern Ireland Environment Agency. All of the weather stations used in the severe weather process have a secondary station which can act as a backup if the primary station fails, or if there are concerns over the accuracy of the data. A station is considered 'frozen' if it records a minimum air temperature below +1°C and a minimum grass temperature below –2°C, and each country is considered to have experienced a 'severe weather day' if more than half of the weather stations in that country experience frozen days. The severe weather process used to include subjective measures of snow cover by meteorological staff working at the weather stations. This is because of the importance of winter stubbles and grasses for wintering geese, and the effect that heavy snow cover can have on depriving geese of these resources. However, these stations are now automatic and so data on snow cover has not been used for some years. The use of satellite imagery was explored but dismissed, primarily due to the costs involved. However, in the last two winters it has become apparent that snow cover may have been affecting temperature readings. This has manifested as recorded grass temperature exceeding the air temperature at a given meteorological station. In coming winters where this occurs then it will be assumed that the thermometer is insulated by snow cover and the data from the secondary (backup) station will be used if available. Severe weather procedure After five consecutive severe weather days the Met Office notifies JNCC, BASC and other key partners (including the Royal Society for the Protection of Birds, the British Trust for Ornithology, the Wildfowl and Wetlands Trust, the Game and Wildlife Conservation Trust and the National Gamekeepers Association). From this point forward the Met Offices notifies the key partners of the previous day’s weather on a daily basis. After seven consecutive severe weather days BASC normally calls on its members to show 'voluntary restraint' where local conditions merit it. This announcement is made on the BASC website, press releases are issued across the UK and individual clubs are emailed directly. There is no obligation or statutory requirement for hunters to take any action at this point, as it is entirely dependent on the local conditions. However, it is normal for hunting clubs in the most severely hit places to put some sort of restrictions in place. These can take the form of reductions in bag size, reductions in the number of visits, voluntary suspensions of all shooting activity, or no action at all. Increasingly efforts are being made to ensure other recreational countryside users such as ramblers and dog walkers are also informed.

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During the period of voluntary restraint, BASC and other key partners ask their members to provide information on the condition of birds and habitats in the local area. There is no obligation for UK hunters to be a member of any hunting organisation. However, of the approximately 170 wildfowling clubs in the UK, 162 are affiliated to BASC so it is therefore likely that BASC is able to communicate with the majority of coastal wildfowlers. The JNCC recording form encourages the submission of information on unusual bird species or numbers, as well as sightings of found dead birds and the state of local water bodies (frozen or not). This information is relayed to JNCC who use it to inform their decision making leading up to, and during, a statutory suspension. After thirteen consecutive severe weather days, the relevant country statutory conservation agency approaches the environment minister to discuss the implementation of a statutory suspension of waterfowl hunting. The relevant agencies are Natural England (NE) in England, Countryside Council for Wales (CCW) in Wales, Scottish Natural Heritage (SNH) in Scotland and the Northern Ireland Environment Agency (NIEA) in Northern Ireland. If the cold weather looks likely to continue then the Protection Order will be signed by the secretary of state and will usually commence two days later. These two days give BASC, and other key partners, time to inform their members of the impending suspension. As with the call for voluntary restraint the BASC website is updated, press releases are issued and clubs are emailed directly. Protection Orders make it illegal to shoot all ducks (including reared, non–wild mallard), geese, coot, moorhen, snipe and woodcock, and are imposed for fourteen days, but are usually reviewed after seven. Game species such as pheasant and partridge are not included in the Protection Order as there is no evidence that continued hunting of these species adversely affects wildfowl (which tend to move to the coast during cold periods) or waders. If the weather conditions improve, the Order may be lifted early, but if it looks like the cold weather will continue the Order will run for its full fourteen days, and potentially a second order can be signed. Following the end of a Protection Order, BASC reviews the extent of the area affected, the length of time the conditions have prevailed for and likely future conditions and may call on its members to exercise voluntary restraint in areas where it is needed. The above process assumes that the severe weather criteria are met every day during the process. In a typical winter this does not happen. Days where half or fewer of the stations in a given country record frozen days are known as 'thaw days'. One or two consecutive thaw days have no effect on the severe weather day count. However, the severe weather day count resets to zero on the third consecutive thaw day. Review process It has been usual for there to be a series of reviews in the spring and summer following a winter where the

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Primary station Secondary station

Fig. 1. Map of the United Kingdom showing primary and secondary meteorological stations used in the severe weather process. For the sake of clarity, secondary stations have not been shown in Northern Ireland. The Republic of Ireland is not part of the UK’s severe weather process. Fig. 1. Mapa del Reino Unido mostrando las estaciones meteorológicas primarias y secundarias utilizadas para reconocer condiciones meteorológicas muy adversas.Para evitar confusiones no se han incluido las estaciones secundarias de Irlanda del Norte. La República de Irlanda no forma parte de estos procedimientos del Reino Unido.

severe weather processes were initiated. These reviews act as an adaptive management process which allows the parties involved to learn from experience and revise the system in light of those experiences. To that end, these reviews have been invaluable in refining the processes and criteria, and through the last 30 years they have led to numerous improvements in the system. These include changes in the number and positioning of meteorological stations, involvement of other countryside users (for example bird watchers) and improvements in communication. Discussion The UK system for managing the shooting of waterfowl during periods of severe weather has worked well for over 30 years. One of the more important factors be-

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hind this success is the good communication between statutory and non–statutory agencies, the shooting community and the conservation bodies. A review of the UK severe weather process (Stroud et al., 2006) suggested that the system may become redundant in the future due to warming winters. In the five years since that review was written the severe weather process has been initiated three times, with statutory suspensions coming in to force in Scotland and Northern Ireland in the last two winters. These events illustrate the difficulties and uncertainties associated with climate change and highlight the need for a robust system to manage disturbance of wintering water birds during extreme weather events. Furthermore, they stress the importance of continuing the adaptive management of the system which has worked so well to date. References Burger, J., 1998. Effects of motorboats and personal watercraft on flight behavior over a colony of Common Terns. Condor, 100: 528–534. Burton, N. H. K., Rehfisch, M. M. & Clark, N. A., 2002. Impacts of disturbance from construction work on the densities and feeding behavior of waterbirds using the intertidal mudflats of Cardiff Bay, UK. Environmental Management, 30: 865–871. Camphuysen, C. J., Ens, B. J., Heg, D., Hulscher, J. B., VanderMeer, J. & Smit, C. J., 1996. Oystercatcher Haematopus ostralegus winter mortality in The Netherlands: The effect of severe weather and food supply. Ardea, 84A: 469–492. Clark, J. A., 2002. Effects of severe weather on wintering waders. M. Phil. Thesis, Univ. of East Anglia. Cresswell, W. & Whitfield, D. P., 2008. How starvation risk in Redshanks Tringa totanus results in predation mortality from Sparrowhawks Accipiter nisus. Ibis, 150: 209–218.

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Davidson, N. C. & Evans, P. R., 1982. Mortality of Redshanks and Oystercatchers from Starvation During Severe Weather. Bird Study, 29: 183–188. Gill, J. A., Norris, K. & Sutherland, W. J., 2001. The effects of disturbance on habitat use by black–tailed godwits Limosa limosa. Journal of Applied Ecology, 38: 846–856. Goss–Custard, J. D., Jenyon, R. A., Jones, R. E., Newbery, P. E. & Williams, R. L. B., 1977. Ecology of Wash. 2. Seasonal–Variation in Feeding Conditions of Wading Birds (Charadrii). Journal of Applied Ecology, 14: 701–719. Hilton, G. M., Ruxton, G. D. & Cresswell, W., 1999. Choice of foraging area with respect to predation risk in redshanks: the effects of weather and predator activity. Oikos, 87: 295–302. Stroud, D. A., Harradine, J. P., Shedden, C., Hughes, J., Williams, G., Clark, J. A. & Clark, N. A., 2006. Reducing waterbird mortality in severe cold weather: 25 years of statutory suspensions in Britain. In: Waterbirds around the world: 784–790 (G. C. Boere, C. A. Galbraith & D. A. Stroud, Eds.). The Stationery Office, Edinburgh, UK. Stroud, J. M., 1992. Statutory suspension of wildfowling in severe weather: review of past winter weather and actions. JNCC Report 75. Peterborough, UK, JNCC. Suter, W. & Vaneerden, M. R., 1992. Simultaneous Mass Starvation of Wintering Diving Ducks in Switzerland and the Netherlands – a Wrong Decision in the Right Strategy. Ardea, 80: 229–242. Tamisier, A., Bechet, A., Jarry, G., Lefeuvre, J. C. & Le Maho, Y., 2003. Effects of hunting disturbance on waterbirds. A review of literature. Revue D Ecologie–La Terre Et La Vie, 58: 435–449. Whitfield, D. P., 2003. Predation by Eurasian sparrowhawks produces density–dependent mortality of wintering redshanks. Journal of Animal Ecology, 72: 27–35.

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Visual counts, bioacoustics and RADAR: three methods to study waterfowl prenuptial migration in Southern France T. Beroud, J. Druais, Y. Bay & J. C. Ricci

Beroud, T., Druais, J., Bay, Y. & Ricci, J. C., 2012. Visual counts, bioacoustics and RADAR: three methods to study waterfowl prenuptial migration in Southern France. Animal Biodiversity and Conservation, 35.2: 175–188. Abstract Visual counts, bioacoustics and RADAR: three methods to study waterfowl prenuptial migration in Southern France.— This study comes from four years (2006–2009) of monitoring on two sites during the prenuptial migration. On each site, a monitoring of 24 hours per each 10–day period from the second 10–day period of January (J2), though February (F1–F3) and March (M1–M3), up to the first 10–day period of April (A1). Monitoring was carried out by RADAR (FURUNO FAR2127), associated with nocturnal bioacoustics recordings, and visual censuses on the same areas. The monitoring effort was considerable: visual counts carried out represent 282 counts–sites (n = 262,030 ducks counted), bioacoustics detected 9,573 calls during 814 hours of nocturnal recording and RADAR recorded 67,368 echoes on a set of 2,128 hours of monitoring. Visual counts showed a decline in the number of birds from late January/early February. Two patterns were observed with the nocturnal recordings with a maximum or a minimum of the value of bioacoustics index on F2 and F3, depending on the years. RADAR, the most relevant method for tracking of bird movements at a population level, identified two different abundance peaks using variables 'flight altitude > 400 m' and 'flight direction towards north–east/south–east', considered as characteritics of the prenuptial migration. The first peak was detected during F1 on Site 1 only in 2007 (once every four years) and during F2 on Site 2 only in 2006 (once every four years). A second peak with a higher number of echoes was recorded on M1 (Site 1) and on M2 (Site 2). Although all methods may suffer from different biases, the combination of two new technologies complementary to visual counts provided reliable and updated data for waterfowl migration in the Mediterranean area. Key words: Prenuptial migration, Waterfowl, Counts, Bioacoustics, RADAR. Resumen Conteos visuales, bioacústica y RADAR: tres métodos para estudiar la migración prenupcial de aves acuáticas en el sur de Francia.— Este estudio es el resultado de cuatro años de monitorización (2006–2009) en dos lugares durante la migración prenupcial. En cada uno, se llevó a cabo un seguimiento de 24 horas, durante periodos de 10 días, a lo largo de la segunda década de enero (J2), el mes de febrero (F1–F3), marzo (M1–M3) y la primera década de abril (A1). La migración se monitorizó mediante RADAR (FURUNO FAR2127), asociado con grabaciones bioacústicas nocturnas, y censos visuales en las mismas áreas. El esfuerzo de muestreo fue considerable: los conteos visuales totalizaron 282 conteos–sitios (n = 262.030 patos contados), mediante bioacústica se obtuvierono 9.573 vocalizaciones en 814 horas de grabación nocturna y mediante RADAR se registraron 67.368 ecos durante 2.128 horas de vigilancia. Los censos visuales muestran una disminución del número de aves a finales de enero/principios de febrero. Los registros nocturnos presentan un máximo o mínimo del índice bioacústico en F2 y F3 función del año. El RADAR, el mejor método para estudiar los movimientos de aves a nivel de población, identificó dos picos de abundancia diferentes, utilizando las variables “altura de vuelo > 400m” y “dirección de vuelo hacia noreste/sureste” consideradas como características de la migración prenupcial. El primer pico se detectó en F1 en el Sitio 1 sólo en el 2007 (un año de cada cuatro) y en F2 en el Sitio 2 sólo en el 2006 (un año de cada cuatro). Un segundo pico, de mayor intensidad, se detectó en M1 (Sitio 1) y en M2 (Sitio 2). Aunque todos los métodos considerados pueden tener sesgos, el uso de dos nuevas tecnologías en combinación con los conteos visuales, nos ha permitido obtener datos fiables y actuales sobre la migración de aves acuáticas en el área mediterránea. ISSN: 1578–665X

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Palabras clave: Migración prenupcial, Aves acuáticas, Conteos, Bioacústica, RADAR. Received: 22 XII 11; Conditional acceptance: 10 II 12; Final acceptance: 5 VI 12 Timothée Beroud, Jennifer Druais, Yannick Bay & Jean–Claude Ricci, Inst. Méditerranéen du Patrimoine Cynégétique et Faunistique (IMPCF), Domaine expérimental agri–environnement, Villa 'Les Bouillens', 30310 Vergèze, France. Corresponding author: Timothée Beroud. E–mail: instmed@impcf.fr

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Introduction Bird migration is an annual journey performed between breeding and wintering areas. We distinguish postnuptial migration (after breeding heading wintering sites) and prenuptial migration (before breeding coming back to nesting sites). The spring migration (prenuptial), the aim of this study, is characterized by flight altitudes higher than the autumn migration (postnuptial) (Elkins, 1996; Bruderer, 1997), and is often carried out more quickly (Arzel et al., 2006) to reach nesting sites sooner and to reproduce in the best conditions (Gordo, 2007). In France, a major route of prenuptial migration crosses the Mediterranean arc (MNHN & ONC, 1989; Dubois & Rousseau, 2005) (fig. 1). Migratory waterfowl come from Spain or Africa and head for the north/north–east direction. Some follow secondary routes and head for the east and north–east direction (Laty, 1979; ONFSH, 2004; ORNIS, 2008). The waterfowl movements are very complex, especially between their wintering sites and their breeding sites (MNHN & ONC, 1989). It is often difficult to distinguish between the local movements and the migratory flights. Moreover, according to the European Directive 2009/147/CE, member states should not hunt migratory bird species 'during their period of reproduction or during their return to their rearing grounds'. Therefore, thorough knowledge of the timing of the prenuptial migration is crucial for the conservation of migratory bird species. The chronology of waterfowl prenuptial migration in France is a subject treated by many authors on a national scale but only with methods such as visual counts or ringing (Fouque et al., 1997; ONFSH, 2004; Guillemain et al., 2006). For this reason we selected two recent technologies for this study: RADAR and bioacoustics. Thus, this study is innovative because it is the first time these three complementary methods are used simultaneously. The aim of this study, in a context of conservation biology, is not to quantify bird migration but rather to determine the period from which the migration variables used and defined in this paper are fulfilled, in the French Mediterranean area. Material and methods This study was carried out on two sites in Southern France, located on the main axis of waterfowl spring migration (fig. 1). The RADAR was located at Fleury d'Aude (Site 1) (43° 12.301 N, 3° 11.525 E), within a wetland situated between the Mediterranean sea and a large scrubland associated to wine–producing areas, near the 'Pissevache pond', not previously reported as a national or international wintering site for ducks (Deceuninck & Fouque, 2010). As for Site 2, the RADAR was located at Saintes Maries de la Mer in the Camargue (43° 29.842 N, 4° 27.032 E), between two main water bodies ('Consécanière pond' and 'Imperial pond'), three kilometers to the north of the Mediterranean sea, within a wetland of international interest for bird conservation, and the most important wintering area for ducks in the French Mediterranean area (Deceuninck & Fouque, 2010).

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125

250

500 km

France

Spain Site 1

Site 2

Fig. 1. Migratory corridors of prenuptial migration in France with the two study sites (modified from Laty in MNHN/ONC, 1989). Fig. 1. Corredores de migración prenupcial en Francia con los dos sitios de estudio (modificado de Laty en MNHN/ONC, 1989).

The monitoring was conducted on each site, and for each 10–day period, from the second 10–day period of January to the first 10–day period of April, from 2006 to 2009. Thus, in this paper, each 10–day period of each month is named with the first letter of the month considered and the number of the corresponding 10–day period (e.g. J2 for the second 10– day period in January). During each 10–day period, RADAR tracking took place for 24 hours, from noon to noon, and the bioacoustics station worked from 8 p.m. to 8 a.m. The visual counts were carried out each morning following RADAR monitoring on various ponds within the same area by one or two observers with a telescope (fig. 2). Counting sites were chosen in order to reflect the best typical duck habitats, and the survey tried to cover all the open water of the pond considered. On Site 1, four ponds were counted on each 10–day period (except for A1 in 2007) during the four years of the study: 'Estagnol pond', 80 ha, two counting points; 'Vic la Gardiole pond', 1,255 ha, but only 520 ha counted with five points; 'la Castillone pond', 75 ha, one counting point and 'Saint Marcel pond', 38 ha, two counting points. In 2008 and 2009, an additional pond was counted: 'Vendres pond',

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1,800 ha, two counting points. Near Site 2, three ponds were counted: 'Consécanière pond', 1,700 ha, two counting points; 'Mas de Tamaris pond', 53 ha, one counting point, and 'Mas de l'Ange pond'; 20 ha, one counting point. The large 'Consécanière pond' is a protected area with no hunting but it is surrounded by many ponds where hunting is allowed, particularly at southern, western and northern points. Hunting was allowed in the remaining wetlands of our sample, such as 'Mas de l'Ange' and 'Mas de Tamaris' ponds. Not all counting points were sampled in every 10–day period. Thus, 'Consécanière pond' was not counted on J2 and A1 in 2006 or F2 in 2007. 'Mas de Tamaris pond' could not be counted on J2, F1 and A1 in 2007 and was dried up in M3 and A1 in 2007 and since M1 in 2008. Similarly, 'Mas de l'Ange pond' was not counted on J2 and A1 in 2006 and was dried up since M2 in 2007 and since F2 in 2008. All these areas are known to be wintering sites (Deceuninck & Fouque, 2010) for ducks but the number of wintering birds is much lower on the Site 1 than in the Camargue (Site 2). Others censuses were performed on two complementary sites ('Canet pond', 1,000 ha, 12 counting points and 'Villeneuve de la Raho pond', 225 ha, 10 counting points) located in the Eastern Pyrenees from 2006 to 2009. These counting sites consist of isolated large ponds, and are known to be wintering sites for ducks (Deceuninck & Fouque, 2010). All counting points at each site were the same throughout the four years of the study. All waterfowl species were counted, but this study focused only on following Anatidae species: mallard (Anas platyrhynchos), teal (Anas crecca), gadwall (Anas strepera), wigeon (Anas penelope), pintail (Anas acuta), garganey (Anas querquedula), shoveller (Anas clypeata), common shelduck (Tadorna tadorna), red–crested pochard (Netta rufina), common pochard (Aythya ferina), tufted duck (Aythya fuligula), and greater scaup (Aythya marila). Bioacoustics recordings were performed on RADAR locations using a bioacoustics station consisting of a microphone (Telinga pro PIP 4, 40 Hz–18 MHz) with a theoretical limit detection of 1000 meters, a parabolic reflector, and a sound numeric recorder (Sony MiniDisc). Recordings were heard and analyzed by a specialized operator. All calls (with the identification of the species when it was possible) and bird movements (such as a characteristic wing noise of a duck) were noted with the time of recording. Only species mentioned previously were considered in this work. Finally, we used a maritime RADAR (FURUNO FAR–2127 BlackBox (X–Band, 9,410 ± 30 MHz), 25 kW power) with an antenna (XN–24AF) of 2.40 m length. The antenna can be elevated to a height of 12 m using a hydraulic platform for a better detection of the targets. The RADAR is connected to a mobile laboratory (camping–car) fitted out with a control screen, a GPS, and a console to adjust the RADAR settings (Seaman, Gain, Range, etc.). Data were recorded using the software RecordRADAR, v1.2 (Pégase Instrumentation) in a computer. The reflectivity of RADAR waves, on the water surface, is limited, or even absent. For this reason the RADAR was located near a large pond on both sites

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in order to avoid the ground clutter. Thus, the echo reading was facilitated throughout the reading line (horizontal position). In vertical position, targets were detected under an altitude of 100 m and over 3,000 m. However, a blind sector of about 50 m around the RADAR, in both positions, prevented any echo reading. Two positions were used for each hour of monitoring (10 minutes of recording in each position). Horizontal position detects and tracks flight directions. Thus, the RADAR was positioned facing the migration front (south–west), with a range of 3 km. For the notation, only echoes crossing an imaginary line perpendicular to the migration front (reading line) were considered for the directions (fig. 3). Therefore, the echoes directions were classified according to four classes: south–west/north–west (SWNW), north–west/north–east (NWNE), north–east/south– east (NESE) and south–east/south–west (SESW). Vertically, the RADAR detects flight altitudes, with a range of 1.5 km, and all the echoes were noted according to two classes that were determined 'a posteriori' after analysing data: below and above 400 m (< 400 m and > 400 m; see below and Results for a justification of this threshold). For each monitoring, RADAR settings were optimized (pulse s2, Gain = 60, Seaman ≥ 20) to detect, with a range between 1.5 km and 3 km, at least mid–sized birds. Overall sampling effort was important: visual counts carried out represent 282 counts–sites (n = 262,030 birds counted), bioacoustics detected 9,573 contacts during 814 hours of nocturnal recording, and RADAR recorded 67,368 echoes on a set of 2,128 hours of monitoring. As explained in the previous section, all ponds may not have been counted during every monitoring for several reasons (weather, pond dried up during the study, etc.). For this reason, the visual count results (analysed with STATISTICA 7.1 software) for each site are presented as the average number of Anatidae per pond counted with the associated standard deviation (fig. 4). Concerning the bioacoustics results, the number of contacts (calls or typical wing noises of ducks) was analyzed per workable hour according to 10–day periods for each year, called 'Bioacoustics index'. Mallards (Anas platyrhynchos) and common shelducks (Tadorna tadorna) were removed from the analysis because these species are mainly sedentary in the study areas (fig. 5, STATISTICA 7.1 software). The data were compared with a random distribution of the bioacoustics index for the nine 10–day periods studied (c² test). Bonferroni intervals (Neu et al., 1974; Byers et al., 1984) were then calculated to identify 10–day periods that differed significantly from the others. Anatidae species are mainly nocturnal migrants (Cramp & Simmons, 1977). Many studies performed in other countries have shown that migration starts about one hour after sunset and reaches a peak between the first and fourth hour of the night (Gauthreaux, 1971; Alerstam, 1976; Richardson, 1978; Laty, 1979; Bruderer, 1997; Zehnder et al., 2002; Bruderer, 2003). So as to avoid the local movement of ducks between their resting areas and their foraging areas (during

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France Site 2

Site 1

Mediterranean Sea

N Eastern Pyrenees

RADAR position

100 km

Counting site

Fig. 2. Study areas with RADAR locations and counting sites. Fig. 2. Zonas de estudio con las posiciones del RADAR y de los lugares de conteo.

the beginning and the end of the day) (Tamisier & Dehorter, 1999), only nocturnal values between 8 pm and 5 am were considered (both for bioacoustics and RADAR results). RADAR results are first presented using the Migration Traffic Rate (MTR) (figs. 6, 7; STATISTICA 7.1 software), widely used in others studies (Lowery, 1951; Bruderer, 1971; Rivera & Bruderer, 1998; LPO/Biotope, 2008; Schmaljohann et al., 2008). MTR is defined as the number of echoes crossing a virtual line of fixed length (1 km) perpendicular to the flight direction within one hour. Therefore, the data recorded during ten minutes (for each antenna position) were converted to obtain a number of echoes per kilometer and per hour. A Principal Component Analysis (PCA) was then performed for data from each site (STATISTICA 7.1 software) in order to compare all the 10–day periods according to nocturnal flight direction and altitude variables (fig. 8). Finally, it is widely accepted that the main flight direction of migrant birds, during prenuptial migration in the study areas is towards the east and north east (Laty, 1979; ONFSH, 2004; ORNIS, 2008). Moreover, nocturnal migratory birds are known to fly at high altitudes during their migration (Kerlinger, 1995; Miller et al., 2005), and according to Newton (2008), the birds fly at low altitudes when they perform local movements. Thus, we assumed that the flight altitudes above 400 m could reflect migrant birds. We therefore considered two variables as characteristics of the prenuptial migration in this study: flight directions towards north–east/ south–east (NESE) and flight altitudes above 400 m (> 400 m).

Results Visual counts On Site 1 (fig. 4), following a decrease in the number of birds until J3, the number of ducks seemed to increase from F1 (2008) or F2 (2006) to M1. In 2007

N NE

SE S

RADAR orientation

Reading line

Fig. 3. Schematic illustration of the RADAR orientation and the reading line for the echoes notation in horizontal position. Fig. 3. Ilustración esquemática de la orientación del RADAR y la línea de la lectura para la notación de los ecos en posición horizontal.

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Site 1 (N = 44,915) Average number of Anatidae/pond counted

1,000 800 600 400 200

0 J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007 1,000 800 600 400 200 0 J2 J3 F1

F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2008 Year 2009 10窶電ay periods

Site 2 (N = 183,904)

Average number of Anatidae/pond counted

1,000 800 600 400 200

0 J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007

1,000 800 600 400 200

0 J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 Year 2008 10窶電ay periods

F2 F3 M1 Year 2009

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Average number of Anatidae/pond counted

Eastern Pyrenees (N = 33,211) 1,400 1,200 1,000 800 600 400 200 0

J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007

1,400 1,200 1,000 800 600 400 200 0

J2 J3

F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2008 Year 2009 10–day periods

Fig. 4. Average number (± SD) of Anatidae per pond counted on Site 1, Site 2 and the Eastern Pyrenees sites, according to 10–day periods and years. Fig. 4. Número medio (± DE) de Anatidae por estanque contados en el Sitio 1, el Sitio 2 y en los sitios de los Pireneos Orientales, según los periodos de diez días y los años.

and 2009, the number of ducks seemed to be relatively equivalent until M1. A marked drop was observed for each year of the study on A1. On Site 2 (fig. 4), a similar trend was observed in 2006 and in 2008, with a decrease in bird numbers since J2 until M2. In 2007, the number of ducks seems to be similar from J3 to F1. An increase was observed on F2 before an important decrease until A1. In 2009, a low decrease in the number of birds was observed from J2 to F1. Then, increases were observed on F2 and on M1. Finally, in the Eastern Pyrenees, a similar trend was observed each year, with a regular decrease in the number of ducks from F1 until A1. However, in 2009, the decrease in the number of birds seems to be observed since J3. Bioacoustics The main species identified by their call were mallards (Anas platyrhynchos), common shelducks (Tadorna tadorna), teals (Anas crecca), gadwalls (Anas strepera), wigeons (Anas penelope), garganeys (Anas querquedula), and shovellers (Anas clypeata).

However, most contacts were represented by bird movements (typical wing noise of ducks). Very few contacts, mallards and common shelducks removed, were recorded on Site 1 (n = 35) as compared to Site 2 (n = 441) (fig. 5). Two patterns were observed. In 2006 and 2008, we recorded an increase until F2, and then a decrease until A1. On the other hand, the opposite was observed in 2007 and 2009, with a decrease until F3, followed by an increase during the month of March. However, the statistical tests performed (c² test and Bonferroni intervals) were not significant (p > 0.05). RADAR On Site 1 (fig. 6), according to the variable 'NESE', the earliest peak of low intensity (11.3 echoes/km/h) was detected during F1, in 2008 only, associated with prevailing flight directions towards NWNE. At no time did NESE flight direction seem to be the prevailing flight direction on F2, but it appeared to be so during F3, only in 2006. On the other hand, NESE flight direction seemed to be considerably favoured

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Bioacoustics index (number of contacts/hour)

Site 1 (N = 35)

Site 2 (N = 441)

6 5 4 3 2 1

0 J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007 6 5 4 3 2 1

0 J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2008 Year 2009 10–day periods Fig. 5. Bioacoustics index (number of contacts/hour) of Anatidae without mallards and common shelducks according to 10–day periods for each year. Fig. 5. Índice de bioacústica de Anatidae sin ánades azulones ni tarros blancos según los periodos de diez días para cada año.

during M1 and M3 throughout the four years of this study (MTR maximum of 57.3 echoes/km/h on M3 in 2008). Regarding the variable '> 400 m', very few echoes were recorded during J2 and J3 (maximum of 6.6 echoes/km/h on J2 in 2006). Then, a first peak of low intensity on F1 (2007) or F2 (2006) was observed. MTR values were always very low on F3 in the four study years. On the other hand, higher MTR values were recorded during M1 in 2006 and 2007, with a maximum of 406.4 echoes/km/h (2007) and during M2 (2006, 2008 and 2009). As for the variable '< 400 m', the MTR values were very low throughout the study period (maximum of 30 echoes/ km/h on M2 in 2008). On Site 2 (fig. 7), regarding the flight direction variables, a background noise was observed in January (which was not detected on Site 1). The variable 'NESE' seems to be privileged, firstly on F2 (2008), with a low MTR value (16 echoes/km/h). Then, this flight direction was considerably favoured on F3 but only in 2008. The most regular movements towards NESE seemed to occur during M2, with non–negligible MTR values in 2006, 2008 and 2009. Very few echoes were recorded, all flight directions combined, throughout the study period in 2007. According to the variable '> 400 m', the earliest peaks were recorded

on F2 in 2006 and 2007, and on F3 in 2007 and 2008. However, the MTR values of flight directions above 400 m were non–negligible on M2 (for the four study years), with the most important peak in 2008 (207.6 echoes/km/h). As for the variable '< 400 m', the MTR values seemed to be relatively low and seemed to increase from the month of March. However, these values were higher than those of Site 1. In order to complete these observations, we performed a multivariate analysis (PCA). On Site 1 (fig. 8A), the two first axes (Fact.1 x Fact.2) of the PCA explained 84.75% of the variance (c²obs = 38.31, p = 8.11 x 10–4). Fact.1 is mainly represented by the flight directions towards North (NESE, 19.85% and NWNE, 14.61%) and by the flight altitudes (< 400 m, 22.75%, and > 400 m, 22.88%), that is to say, 80%. The direction towards SESW contributed for 19.85%. As for Fact.2, this axis was mainly represented by flight directions towards SWNW (74.6%) and towards SESW (12.68%), that is to say a total of 86.74%. This method separated two groups, opposed when we consider the Fact.1. The first group consists of 10–day periods in January and February and the second group of those in March, characterized by flight altitude variables and flight directions towards NESE and towards NWSE.

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SESW direction (N = 493)

NWNE direction (N = 1,166)

SWNW direction (N = 291)

NESE direction (N = 2,836)

60 50 40 Directional MTR (number of echoes/km/h)

30 20 10

0 J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007 60 50 40 30 20 10 0 J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2008 Year 2009 10–day period

Altitudinal MTR (number of echoes/km/h)

Altitude < 400 m (N = 1,296)

Altitude > 400 m (N = 6,515)

450 400 350 300 250 200 150 100 50 0

J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007 450 400 350 300 250 200 150 100 50 0

J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2008 Year 2009 10–day period

Fig. 6. Evolution of nocturnal MTR according to 10–day periods and years on Site 1. Fig. 6. Evolución de las MTR nocturnas, según los periodos de diez días y los años en el Sitio 1.

Directional MTR (number of echoes/km/h)

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SESW direction (N = 720)

NWNE direction (N = 1,149)

SWNW direction (N = 730)

NESE direction (N = 2,047)

70 60 50 40 30 20 10 0

J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007 70 60 50 40 30 20 10 0

J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2008 Year 2009 10–day period

Altitudinal MTR (number of echoes/km/h)

Altitude < 400 m (N = 3,544)

Altitude > 400 m (N = 4,968)

220 200 180 160 140 120 100 80 60 40 20 0

J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2006 Year 2007 220 200 180 160 140 120 100 80 60 40 20 0

J2 J3 F1 F2 F3 M1 M2 M3 A1 J2 J3 F1 F2 F3 M1 M2 M3 A1 Year 2008 Year 2009 10–day period Fig. 7. Evolution of nocturnal MTR according to 10–day periods and years on Site 2. Fig. 7. Evolución de las MTR nocturnas, según los periodos de diez días y los años en el Sitio 2.

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A SWNW

0.5

SESW

Fact.2: 21.29%

Fact 2: 21.29%

1.0

> 400 m < 400 m

NWNE NESE

–0.5

–1.0 –1.0

–0.5

0 0.5 Fact.1: 63.55%

3.5 3.0 2.5 20. 1.5 1.0 0.5 0 –0.5 –1.0 –1.5 –2.0 –2.5 –3.0 –5

F2 F3

–4

–3

–2

–1 0 1 Fact.1: 63.55%

4 3

0.5

Fact.2: 29.95%

SWNW SESW

1.0

> 400 m < 400 m NWNE NESE 0

–0.5

2 F2

1 0

–1

J3 M1 F1

–2

–3 –4

–1.0 –1.0

–0.5 0 0.5 Fact.1: 55.60%

1.0

–5 –6

–5

–4

–3

–2 –1 0 1 Fact.1: 55.60%

Fig. 8. Principal Component Analysis of average nocturnal MTR (2006 to 2009): A. Site 1, Feury d'Aude; B. Site 2, Camargue; J2, J3, F1, F2, F3, M1, M2, M3, A1. 10–day periods; NESE. Average nocturnal MTR towards North–East/South East; NWNE. Average nocturnal MTR towards North West/North East; SESW. Average nocturnal MTR towards South East/South West; SWNW. Average nocturnal MTR towards South West/North West; < 400 m. Average nocturnal MTR of flight altitude below 400 m; > 400 m. Average nocturnal MTR of flight altitude above 400 m. Fig. 8. Análisis de Componentes Principales de los promedios de las MTR (tasas de tráfico migratorio) nocturnas (2006–2009): A. Sitio 1, Feury d'Aude; B. Sitio 2, Camargue; J2, J3, F1, F2, F3, M1, M2, M3, A1. Décadas; NESE. MTR nocturna media hacia el noreste/sureste; NWNE. MTR nocturna media hacia el noroeste/noreste; SESW. MTR nocturna media hacia el sureste/suroeste; SWNW. MTR nocturna media hacia el suroeste/noroeste; < 400 m. MTR nocturna media de las alturas de vuelo por debajo de 400 m; > 400 m. MTR nocturna media de las alturas de vuelo por encima de los 400 m.

On Site 2 (fig. 8B), the Fact.1 and Fact.2 axes of the PCA explained 85.5% of the variance (c²obs = 45.59, p = 6.17 x 10–5). Fact.1 was mainly represented by the flight directions towards north (NESE, 26.85% and NWNE, 22.30%) and by the flight altitudes (< 400 m, 24.82%, and > 400 m, 21.19%), that is to say, more than 95%. As for Fact.2, this axis is mainly repre-

sented by flight directions towards SWNW (47.73%) and towards SESW (46.20%), that is to say, a total of 93.93%. This method also separated two groups of 10–day periods (opposed on Fact.1) with one gathering F3 and M2, characterized by flight direction towards NESE and towards NWNE and flight altitude variables.

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Discussion Our results support that the use of two recently developed technologies (RADAR and bioacoustics recording) combined with a conventional method such as visual counts may improve the study of migratory bird movements. Visual counts provided an instantaneous number estimated throughout the study period. Thus, the analysis focused on bird number variations, but it is difficult to know what kind of movement may explain changes in numbers (migratory or local). This 'classical' approach seemed to identify two different patterns depending on the study areas. A regular decrease in duck numbers was observed on Site 2 and the Eastern Pyrenees since early February. According to MNHN & ONC (1989), this is considered typical of wintering sites and could be interpreted as wintering birds leaving their wintering areas. However, the interpretation of these trends is still difficult. A decrease in the number of birds from January to April can be interpreted as the result of local migratory birds leaving wintering areas, but it may also be the result of local movements due to several factors (disturbance, bad weather, food availability, etc.) (MNHN/ONC, 1989; Fouque et al., 1997; Guillemain et al., 2006). Moreover, decreases in numbers observed until late January at the study sites (fig. 4) could also be due to hunting, authorized until January 31st in France, a possibility we can not evaluate because hunting bags were unknown. The interpretation of increases in the number of birds observed during the study period is also difficult because this method does not give any information about flight directions or flight altitudes. Nevertheless, these observations may be interesting on Site 1, considered as a stopover area. In this case, an increase from mid–February through March (detected in three of four years) would better reflect migration, and could be interpreted as a consequence of stopover behaviour. Furthermore, we can not discard that data could be flawed by an observer effect, because depending on the years and the sites, observers may have changed. Finally, with only three censuses per month, the number of birds may remain stable if the number of incoming birds compensates for the number of departing birds (MNHN/ONC, 1989; Fouque et al., 1997). Regarding bioacoustic results, this method seems to highlight F2 and F3 as turning point periods, because depending on the years, we observed the maximum (F2 in 2006 and 2008) or minimum (F3 in 2007 and 2009) values. However, it is difficult to reach sound conclusions because no tests performed were significant. Ricci et al. (1995) have worked successfully on thrushes, because these species emit many calls during their nocturnal migration flights. This is not the case with ducks. The analysis was therefore mainly based on the wing noise characteristic of ducks, minimizing the number of contacts. Moreover, such as with the first method, bioacoustic recordings can not detect the flight directions. Thus, the bioacoustics method does not seem to be appropriate to study this kind of species. As we have seen, different biases may be linked to all methods. Overall, RADAR tracking seems to be the

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most suitable method to study migratory duck movements at a population level. This tool is particularly interesting to study nocturnal movements (Bruderer, 2003) because it can detect targets to over 3 km, and it overcomes weather conditions (except for extreme events). We also know the flight directions and the flight altitudes, two variables needed to study migratory bird movements (at the strict sense). However, the RADAR used did not allow a wing beat analysis. The main current limitation is therefore the difficulty to identify echoes of birds at the species level (Hamer Environnemental, 2008; Schmaljohann et al., 2008). However, RADAR parameters were settled to avoid the detection of small–sized targets. Moreover, working within a range of 3 km (horizontally) and 1.5 km (vertically), and according to Alerstam (1976), we can assume that the echoes recorded exclude insects and small–sized migrants, such as small passerines (warblers, flycatchers, etc.). Furthermore, according to Arzel et al. (2006), Anatidae species are early migrants compared to other birds. The first migration peaks observed may therefore correspond to these species. The study sites were chosen for their location on the Mediterranean axis of the spring migration. Site 1 is particularly interesting because the wintering birds (Anatidae species) are fewer than on Site 2. This difference involves a limited 'background noise', contrary to Site 2 (fig. 7), due to an important number of local wintering birds. In future, in order to complete this study, it would be interesting to work on a site located at the north of Site 2 (Rhône valley, for example), to study the bird movements leaving the Camargue wintering area. The absence of background noise (wintering) would facilitate the interpretation of data, as was done by Ricci et al. (1995) regarding the chronology of the prenuptial migration of thrushes, with the simultaneous use of bioacoustic monitioring stations in areas with no wintering. However, in this case, it would remain difficult to distinguish birds leaving the Camargue (local departure of wintering birds) from those who fly over without stopping. This study is based on two variables we considered as characteristics of migration: 'NESE' and '> 400 m'. Considering the MTR (figs. 6, 7), very low values were obtained for the variable '< 400 m' throughout the study on Site 1. On the other hand, the MTR values of the variable '< 400 m' recorded on Site 2 were slightly higher, particularly from F3 and during the month of March. We can not exclude that these echoes may correspond to migrant birds decreasing their flight altitude in preparation for stopover, or birds leaving their wintering area. However, from J2 to F2, these flows < 400 m were not associated with a prevailing flight direction. Although the distinction between migratory movements and more local movements can be arbitrary, and despite the fact that migratory movements with flight altitude < 400 m are possible, the results obtained, and particularly on Site 1 with a limited wintering population, would support that flight altitudes > 400 m, associated to flight directions towards north–east/south–east, could be used to study prenuptial migration. According to the PCA (figs. 8A, 8B), the 10–day periods of January do not seem to be

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explained by the characteristic variables used (Fact.1). It appears there were no migratory movements during this period. Although two small increases were observed with the directional and the altitudinal MTR on F1 and F2 (figs. 6, 7), the PCA include these 10–day periods with those of January, indicating that these two 10–day periods are not linked as a migratory period. Moreover, the first peak detected during F1 on the Site 1 (fig. 6) is due to the winter 2007, particularly clement. Indeed, January and February 2007 were among the hottest months for the period 1950–2007, with a temperature of more than 3 degrees Celsius higher than the average temperature (Réseau OEZH, 2008). A peak was detected on F2 in 2006 (once every four years), but only for the variable '> 400 m'. Guillemain et al. (2006) have shown, by ringing, that teal (Anas crecca) spring migration in the Camargue (Site 2) starts during F1. Our results would not be in accordance with this conclusion, because the earliest movements (of low intensity) were detected during F2 in the same area (Camargue, Site 2). As Guillemain et al. (2006) suggest in their paper (data are from 1952 to 1978), migration dates may have changed due to the climate change or the modification of duck habitats. Therefore, these flows recorded during F2 could correspond to birds moving over short distances (possibility of leaving their wintering area), erratically, towards other feeding sites in order to obtain sufficient energy (as fats) to prepare their migration (Fouque et al., 1997; Newton, 2008). It may also be due to the phenomenon of nomadism (Newton, 2008; Boere & Dodman, 2010). Only an intensive monitoring on a large number of birds by Argos transmitters would allow these hypotheses to be tested. The MTR values and the PCA analysis, particularly at Site 2, seem to show that the first migratory movements occur during F3. We may assume that this 10–day period is a turning point in the prenuptial migration, separating two periods of different flow intensities. Indeed, the migration peaks with high values for the two migratory variables were detected in March (M1 on the Site 1 [fig. 6], and M2 on the Site 2 [fig. 7]). These results are consistent with those obtained by RADAR at a national level (LPO/Biotope, 2008), estimating the period of prenuptial migration (all bird species combined) between March and mid–May. During M1 and M2, the RADAR echoes correspond to a wide diversity of nocturnal migratory bird species, including the Anatidae species, which are early migrants compared to many other birds (Arzel et al., 2006). Although visual counts show a first decline in the number of birds from late January/early February, RADAR does not detect significant migratory movements before F3. Contrary to Guillemain et al. (2006), this study does not allow us to assert that diurnal censuses are a reliable method to study migration, particularly the complex prenuptial migration that is well defined in ORNIS (2008). This may be particularly true in an area like southern France, where sedentary and wintering populations may mix with migrant ducks coming from more southern locations. However, this study has shown that results provided by visual counts at the site considered as stopover (Site 1) were more similar to

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RADAR results than those of an important wintering site (Site 2). Therefore, in order to study prenuptial migration in the future, we suggest that visual counts may be a feasible technique, but only at sites that are important as a stopover. We could have focused on one site but it seemed more appropriate to work on two remote sites to obtain more consistent results across the Mediterranean area. This paper has also demonstrated the need for RADAR use as the only way to study flight altitude and flight direction of nocturnal migratory birds at the population level. Although this tool is less precise for tracking only one species (unlike Argos transmitters or ringing), it is the most reliable for multi–species data, particularly for nocturnal migrants. In order to increase tracking efficiency, it would be interesting to use several Radars simultaneously on various sites with no wintering, but the cost in equipment and staff would be considerable. However, in this study, the results obtained provide scientifically relevant data about the phenology of prenuptial migration of duck species sharing a common area like wetlands, which is consistent with the principle of conservation biology. To conclude, this article shows that the most significant period of prenuptial migration in the studied areas started during the first 10–day period of March. In most years we also found evidence of less important migratory movements during the third decade of February, probably reflecting inter–annual and inter–species variability. Finally, this study shows that there is no high–altitude NESE–directed migration up to F2, except for an atypically mild winter like 2007 (which was one of the hottest since the 1950s), when migratory movements could also be assumed in F1. To confirm these hypotheses, more research is needed and it will be interesting, in the future to performing RADAR monitoring at the north of Camargue, on a site with no wintering, to track ducks leaving this wintering area. Acknowledgements We would like to thank all colleagues who contributed to this study, especially technical services of the Hunters Departmental Federations (11, 13, 34, 66 and 84). We are also grateful for financial support from the Hunters National Federation (FNC) and the Hunters Regional Federation of Provence Alpes Côte d'Azur (FRC–PACA). Finally, we would like to thank three anonymous referees for their valuable comments on the first version of the manuscript. References Alerstam, T., 1976. Nocturnal migration of thrushes (Turdus spp.) in Southern Sweden. Oikos, 27: 457–475. Arzel, C., Elmberg, J. & Guillemain, M., 2006. Ecology of spring–migrating Anatidae: a review. Journal of Ornithology, 147: 167–184. Boere, G. C. & Dodman, T., 2010. The Flyway Approach to the Conservation and Wise Use of Waterbirds and

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Wetlands: A Training Kit. Module 1: Understanding the Flyway Approach to Conservation. Wings Over Wetlands Project, Wetlands International and BirdLife International, Ede, The Netherlands. Bruderer, B., 1971. RADARbeobachtungen über den Frühlingszug im Schweizerischen Mittelland. (Ein Beitrag zum Problem der Witterungsabhängigkeit des Vogelzugs). Ornithologische Beobachter, 68: 89–158. – 1997. The study of bird migration by RADAR: part 2: Major achievements. Naturwissenschaften, 84: 45–54. – 2003. The RADAR window to bird migration. In: Avian migration: 347–358 (P. Berthold, E. Gwinner, E. Sonnenschein, Eds). Springer. Berlin. Byers, C. R., Steinhorst, R. K. & Krausman, P. R., 1984. Clarification of a technique for analysis of utilization–availability data. The Journal of Wildlife Management, 48: 1050–1053. Cramp, S. & Simmons, K. E. L., 1977. Handbook of the Birds of Europe, the Middle East and North Africa; the birds of the Western Palearctic, Vol. 1. Oxford Univ. Press, Oxford. Deceuninck, B. & Fouque, C., 2010. Canards dénombrés en France en hiver: importance des zones humides et tendances. Ornithos, 17: 266–283. Dubois, P. J. & Rousseau, E., 2005. La France à tire–d'aile: Comprendre et observer les migrations d'oiseaux. Delachaux et Niestlé, Paris. Elkins, N., 1996. Les oiseaux et la météo. L'influence du temps sur leur comportement. Delachaux et Niestlé, Paris. Fouque, C., Schricke, V. & Barthe, C., 1997. Analyse de l'enquête sur la migration prénuptiale des anatidés d'après le suivi de 32 sites entre 1992 et 1997: variations du début de migration par espèce et effet d'une vague de froid. Internal report ONC/CNERA AM/Réseau Oiseaux d'eau, Paris. Gauthreaux, S. A., 1971. A RADAR and direct visual study of passerine spring migration in Southern Louisiana. The Auk, 88: 343–365. Gordo, O., 2007. Why are bird migration dates shifting? A review of weather and climate effects on avian migratory phenology. Climate research, 35: 37–58. Guillemain, M., Arzel, C., Mondain–Monval, J. Y., Schricke, V., Johnson, A. R. & Simon, G., 2006. Spring migration dates of teal Anas crecca ringed in the Camargue, Southern France. Wildlife Biology, 12: 163–169. Hamer Environnemental, 2008. Synthèse bibliographique sur l'expérience américaine en matière de RADAR utilisé dans le cadre d'études de l'avifaune. France. Kerlinger, P., 1995. How birds migrate. Stackpole Books, Mechanicsburg, Pennsylvania, USA. Laty, M., 1979. Recherches en cours sur les migrations des oiseaux. Départs de Camargue au printemps et arrivées de migrateurs sur la côte méditerranéenne

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Spatial behaviour and survival of translocated wild brown hares C. Fischer & R. Tagand

Fischer, C. & Tagand, R., 2012. Spatial behaviour and survival of translocated wild brown hares. Animal Biodiversity and Conservation, 35.2: 189–196. Abstract Spatial behaviour and survival of translocated wild brown hares.— The fragility of many populations of brown hares in Western Europe is a concern for managers, hunters and naturalists. We took advantage of a locally high density population to use wild individuals to restock areas where the species had disappeared or was close to disappearing. The aim of the project was to assess the evolution of the spatial behaviour after release using radio–tracking. Over 150 wild brown hares were translocated, one third of which were fitted with radio collars. In addition, fifteen individuals were radio–tagged and released back into the source population as a control. Most individuals settled in less than two months and their seasonal home range, once settled, was similar to that observed in the source population. Mean duration of tracking was not significantly different between the two groups. Moreover, two years after the last translocation, tagged individuals can still be observed, but most hares present are not tagged, which indicates natural reproduction of the released individuals. The translocation of wild individuals thus appears to give encouraging results. Key words: Brown hare, Lepus europaeus, Translocation, Home range, Survival, Monitoring. Resumen Conducta espacial y supervivencia de liebres europeas salvajes desplazadas.— La fragilidad de muchas poblaciones de liebres europeas en Europa occidental es una gran preocupación para gestores, cazadores y naturalistas. Aprovechamos la ocasión de disponer de una población local de alta densidad para repoblar con ejemplares salvajes áreas dónde la especie había desaparecido o estaba próxima a la extinción. El objetivo de este proyecto era evaluar la evolución de la conducta espacial tras soltar a los individuos, utilizando el radio–seguimiento. Se trasladaron más de 150 liebres europeas salvajes, a un tercio de las cuales se les puso un collar dotado de un transmisor. Además, quince individuos fueron dotados también con un transmisor, y vueltos a soltar en su población de origen, como grupo de control. La mayoría de los individuos se establecieron en menos de dos meses, y su área de deambulación estacional, una vez establecida, era similar a la observada en la población de origen. La duración media del seguimiento no fue significativamente distinta entre los dos grupos. Dos años después del último traslado, aún pueden ser observados individuos marcados, aunque la mayoría de las liebres presentes en el lugar no están marcadas, lo que indica que ha habido una reproducción natural entre los individuos soltados. Por lo tanto, el desplazamiento de individuos salvajes parece producir resultados esperanzadores. Palabras clave: Liebre europea, Lepus europaeus, Desplazamiento, Área de deambulación, Supervivencia, Monitorización. Received: 23 XII 11; Conditional acceptance: 10 II 12; Final acceptance: 25 V 12 Claude Fischer & Romain Tagand, Dept. Nature Management, Univ. of Applied Sciences of Western Switzerland, 150 rte de Presinge, 1254 Jussy, Switzerland. Corresponding author: Claude Fischer. E–mail: claude.fischer@hesge.ch ISSN: 1578–665X

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Introduction The fragility of many populations of brown hares (Lepus europaeus) in Western Europe is a concern for managers, hunters and naturalists. The influence of multiple factors on the survival of its populations and a habitat under strong competition for space renders conservation efforts particularly difficult. The intensification of agriculture is considered to be a major factor in the decline of hare populations as it leads to a loss of habitat heterogeneity, a reduction of permanent vegetation cover, an increase in the use of pesticides, and precocial mowing of pastures (Hackländer et al., 2002; Rühe & Hohmann, 2004; Baldi & Farago, 2007; Fischer, 2010; Reid et al., 2010). Improving the quality of the habitats is, however, often not enough for the populations to recover (Zellweger–Fischer et al., 2011), and restocking efforts are a widespread additional management tool (Stamatis et al., 2007). However, the effects of this latter method on the survival of the released individuals and on the dynamics of the local hare population are considered and analysed in few instances. In Switzerland, hare populations have experienced a strong decline since the 1960s as in many areas in Western Europe (Vaughan et al., 2003; Smith et al., 2005). Mean hare density in Switzerland is estimated to be under 3 individuals per 100 ha (Zellweger–Fischer, 2011). Canton Geneva has been an exception the last twenty years, as mean populations were always much higher than the nation’s mean, with estimated densities around 12 to 15 individuals per 100 ha. Since 2001, one of the populations in the east of the canton even showed a dramatic increase, reaching a density of over 50 individuals per 100 ha on an area of about 20 km2 (Nature and Landscape Office, administration of Geneva, unpublished data). During these years, damage to crops due to hares also increased markedly. This situation led to two contrasting management needs between neighbouring areas. On the one hand, in Geneva the apparent link between increasing densities and the increase of damage on crops (mainly sunflower, soybean, and peas) led to the willingness to reduce hare populations. The option of culling part of the population was soon abandoned however because of the role that the canton plays regionally regarding the conservation of hare populations, because of the emblematic status of the species for the public, and last but not least, because hunting has been abolished in Geneva since 1974. On the other hand, in many surrounding areas, in Switzerland and in France, the interests of local managers are the exact opposite. There, efforts are made to restore or reintroduce populations. Considering these contrasting management needs, we took advantage of the presence of the locally high density population in Geneva to use wild individuals for restocking experiences in areas were the species had disappeared or was close to disappearing. This management tool, translocation, is increasingly used to restore endangered populations, particularly birds and mammals (Seddon et al., 2007). Regarding hares, however, it is rather unusual. In France, translocation

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was used since the 1970s either with captive bred leverets (Marboutin et al., 1990) or with wild individuals originating from Central or Eastern Europe (Benmergui et al., 1990). There was thus no consideration for the genetic homogeneity of the released hares. In Italy, translocation experiments with wild hares were conducted on a regional scale (Paci et al., 2006; Pelorosso et al., 2008) , respecting the genetic origin of the local population. However, in these instances results were generally disappointing. Translocation in general is often considered as being unsuccessful as released individuals fail to establish viable populations (Angelici et al., 2000; Teixeira et al., 2007). The main factors explaining the poor success of translocations are prolonged stress during and particularly after translocation when released individuals have to adapt to a novel environment, as well as the increased risks they are exposed to during the first days after release when they move around in search of food resources and cover. In our study we tried to reduce stress to a maximum by reducing the time of handling and transport. The aim of the project for the areas to be restocked was thus: (1) to assess the evolution of spatial behaviour after release using radio–tracking, and to compare the results with control individuals tagged in the source population; (2) to measure the sustainability of the method in assessing the survival and reproduction potential of the released individuals; and (3) to compare the results with the frequently used release of captive bred individuals, the main hypothesis being that wild individuals would show a better survival after release than captive bred individuals. Methods Study areas The population where hares were captured, the source population, is located in the South–East of canton Geneva. It is dominated by intensive cultivation of cereal crops interspersed by small woods and fallow strips. A large forest constitutes a limit to the north and east, marking the border with France, Lake Geneva constitutes the limit to the west, and the agglomeration of Geneva is the limit to the South (fig. 1). The area is criss–crossed by an intensive road network and traffic is heavy. Hunting was abolished in the canton in 1974 but is still allowed in the neighbouring areas in France. Hares were released in four different target populations, one in the Canton Valais, Switzerland, and three in Haute–Savoie, France (fig. 1). The target area in the Valais (3000 ha) and two of the target areas in Haute–Savoie (Sciez [1,100 ha] and Arenthon [3,500 ha]) are a mosaic of pastures, crops and woods lying at a similar altitude as the source area (350–450 m). The areas are delimited by roads with heavy traffic and by agglomerations. The last target area in Haute–Savoie, Mieussy (5,200 ha) lies, however, at a much higher altitude (1,500 to 2,000 m). There the landscape is dominated by pastures and prairies, and half of the area is a ski–resort during winter. The area is limited by natural features like high mountain peaks and cliffs.

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Germany

Schafthausen

France Thononles-Bains

Nyon

La Rochesur-Foron Bellagardesur-Valserine

Martigny

Zürichsee

Biel NeuchâtelBern Lucerne ra

u JFribourg

Austria Chur

Thun

Lausanne

Morzine

Geneve

Bodensee

Besel Winterthur Zürich St Gallen

Lake Geneva

Geneva

Lugano

Italy

Sallanches

Annecy Thônes

Fig. 1. Situation of the canton Geneva and of the different study sites: Source population; Target populations. (Switzerland is white and France light grey. Lake Geneva is represented in black.) Fig. 1. Situación del cantón de Ginebra y los distintos lugares de estudio: Población de origen; Poblaciones de destino. (Suiza está en blanco y Francia en gris claro. El lago de Ginebra se ha representado en negro.)

The source population and selected release populations were located in a radius of less than 30 km, except for the Valais distant of 100 km. This allowed reducing time of transport, to release hares of potentially the same genetic origin as the residents, and to release them in areas with similar climatic conditions (with the exception of Mieussy). To get hares from Geneva, managers of the target areas adopted a commitment to stop hunting activities for at least five years, to participate actively in the capture efforts, to improve the habitats, and to play an active role in the monitoring of the restocked populations by conducting spot–light censuses. Captures and marking Most hares were captured using nets after being flushed during silent beats. A few individuals were also captured in cage–traps. Captures were conducted between early November and the end of February in the winters of 2006–07, 2007–08, and 2008–09. Capturing in winter allows beats in agricultural areas without damaging crops and avoids capturing lactating females. Captured individuals were weighed, sexed and ear–tagged. Only those over 3.8 kg were fitted with radio transmitters (TW5SM Biotrack Ltd.) as collars have to be fitted tightly and are thus not suitable for individuals that are not fully grown up. A reflective adhesive was stuck on the ear–tags so they could be seen better during spotlight censuses. All captured hares were then kept in individual transport boxes until release. Some were released back into the source population, directly at the site of capture, as controls. Translocated individuals were released in the various target sites three to five hours after capture.

Captive bred individuals To compare the results of our translocation project with the more usual restocking method using captive bred hares, we took benefit from a trial of a hunting association located in an area bordering one of our target areas. The association brought 22 individuals from a farm in Brittany (France) and released them after a 12–hour train journey. Before release, we fitted eight of them with collars to monitor their spatial use and survival. Monitoring Spotlight censuses were used to determine survival of the released individuals. This method also gives an opportunity to assess population trends and to collect information regarding reproduction success. Even if it does not give precise survival and reproduction rates, the method allows a good insight into the success of the translocation efforts. The number of individuals with tags observed from year to year provides an indication of the survival of the released population, and the proportion of untagged individuals gives an idea on the reproduction success, as all but one target area had no hares left before release and as these areas are largely isolated from potential neighbouring populations. Yearly spotlight censuses have been regularly made for over 20 years. They were, however, more focused on other game species, such as roe deer (Capreolus capreolus), in the target areas as hares had not been seen since the early nineties in Arenthon and Sciez, and were only rarely spotted in Mieussy (Fédération de chasse 74, pers. comm.). Spotlight censuses were then intensified as soon as our project was started and homogenised between the different study areas. Since then, sessions

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are conducted in early March each year and repeated at least 4 times within 3–4 weeks. Spotlight censuses should continue in the future. In this paper we present the results of the last census, in March 2011, that is two years after the most recent release. The evolution of spatial use and the degree of settlement of the translocated hares were assessed by radio–tracking some of the released individuals. To assess the degree of settlement, that is, the spatial stabilisation, we considered the evolution of the mean distance between two consecutive fixes. As proposed in Moreno et al. (2004), a hare was considered stabilised when for three or more consecutive days this distance was less than all of its averaged values. In addition, an idea of the dispersion distance was given by considering the distance between each successive location and release site (Pépin & Cargnelutti, 1985). In a suitable area, released hares should show a stabilisation of their home range and should not remain erratic. Comparing the stabilised home ranges in the released populations to those of the source population gives a rough idea of the habitat quality, although factors such as hare density (Ferretti et al., 2010) or stress (Teixeira et al., 2007) also play a significant role. Each individual was located once at night and once during daytime, every second day. Intensive tracking of 8 hours in a row with fixes taken every 30 minutes was performed on every individual on average every second week depending on the number of tracked individuals at a given moment. To avoid any autocorrelation, we considered only fixes separated by more than two hours for the calculation of home ranges. In the mountain area of Mieussy, half the released individuals established in steep slopes with a South – South–west orientation). The accessibility of these areas was particularly difficult in winter, as the snow cover can reach over one meter, but also during the rest of the year because of the steep slippery slopes. Furthermore, this particular topography made the tracking very difficult because of the echoes and because of the shielding effect of the mountain. In this area it was thus often not possible to follow our tracking protocol. Home ranges were evaluated by the Minimal Convex Polygon (MCP, at 95%), to be comparable to data from the literature, and by the adaptive Kernel method (at 95%, with least–square cross–validation for bandwidth selection). The analyses were conducted using Arcview 9.3 with HRT tools extension. The latter allows to avoid overestimating the area of the home range if location points are clustered (Ryan, 2011) and was thus used for further analysis. Home ranges were only calculated for individuals with more than 50 fixes to avoid an overestimation (Seaman et al., 1999). For individuals that were determined to have settled, home ranges were calculated from the moment of stabilisation. Finally, radio–tracking data also gave an indication of the survival duration of the individuals fitted with transmitters simply by considering the number of days (= radio–days) each tracked individual survived. It allowed a 'mean tracking duration to be determined in each area. Spotlight censuses were conducted in all target areas as well as in the source area, and should be

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continued in the long term. The monitoring through radio–tracking was conducted in two of the target areas (Arenthon and Mieussy) and in the source area, as a control. Results Between 2006 and 2009, a total of 185 wild Brown hares was captured in the source population. Among these, 159 individuals were translocated to the four target areas (table 1). Of these, 58 were fitted with transmitters. In addition, fifteen individuals were radio– tagged and released back into the source population as controls. The eleven remaining individuals died after capture or during transport. Survival For both target areas, the mean tracking duration (= radio–days) of the individuals fitted with transmitters was not significantly different from that of the source population (136 days ± 144 and 173 ± 197 for Arenthon (N = 18) and Mieussy (N = 10) respectively, compared to 165 days ±126 for the source population (N = 13); Mann–Whitney U–test: P > 0.05). Considering spotlight counts, the total number of hares recorded two years after the last release varied between 40 and 55% of the number of translocated individuals (table 2). Ten to 30% of them were tagged, and had thus survived for over two years. Spatial use After release, hares did not show any preference in the directions taken and thus gave no indication of a possible homing behaviour. Only a minority of them went out of the target areas (11.1% in Arenthon, 9.6% in Mieussy), even after several months of tracking. Evolution of the distances between each successive fix and release point usually stabilised at 758 m (± 316) and 20 days in Arenthon, and at 1,428 m (± 1,185) and 50 days in Mieussy. At first sight, home ranges in the target areas were significantly larger than for control individuals from the source population (table 3). However, considering that hares in the target areas took 20 and 50 days, respectively, to settle, we calculated home ranges one and two months after release (table 4). Differences between source and target areas were no longer statistically significant . It should be pointed out that there was a high individual variation in home range sizes in all the study areas. Captive bred individuals Of the 22 captive–bred hares released, none were re– observed during spot–light censuses conducted nine months later. Eight of them were fitted with collars, and seven of these died within the first 36 hours after release; the last one survived 20 days. Translated into radio–day, captive–bred hares thus survived for a mean of 3.4 (± 6.5) days, meaning significantly less

Animal Biodiversity and Conservation 35.2 (2012)

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Table 1. Number of hares released in the different study areas and of individual fitted with transmitters: N. Total number released; F. Females; M. Males. Tabla 1. Número de liebres liberadas en las distintas áreas de estudio y de individuos equipados con transmisores: N. Número total de individuos liberados; M. Machos; F. Hembras. Study area

With transmitters

> 50 fixes

Source (Geneva)

Arenthon

Mieussy

Sciez

–

Valais

–

Total

174

Table 2. Results of spotlight counts: * Individuals withdrawn from the population (159 translocated + 11 dead); (1) Two years after last release; Tabla 2. Resultado de los censos con focos: * Individuos sacados de la población (159 desplazados + 11 muertos); (1) Dos años después de la última liberación. Area Initial situation

Number released 2006–09

Re–obs.(1)

Without tags(1)

160–210

–170*

None

140–180

Arenthon

None

Mieussy

< 3

Source

Total obs.(1)

Sciez

None

Valais

None

Table 3. Mean home range sizes as calculated by the Minimum Convex Polygon (MCP) and the Kernel method. (Standard deviation is given in brackets.) Tabla 3. Tamaño medio de cada área de deambulación calculados según el Polígono Convexo Mínimo (MCP) y el método del núcleo (método Kernel). (Desviación estándar entre paréntesis.) Area Source (N = 13) Arenthon (N = 18) Mieussy (N = 8)

MCP 95%

MCP 50%

Kernel 95%

Kernel 50%

31.8 ha (± 18.8)

7.8 ha (± 6.8)

39.1 ha (± 24.9)

6 ha (± 4.8)

162.4 ha (± 196.7)

25 ha (± 14.3)

118.4 ha (± 113)

16 ha (± 11)

117.2 ha (± 75.6)

15.4 ha (± 8.8)

77.5 ha (± 39.8)

9.3 ha (± 5.1)

than hares tracked in the source population (Mann– Withney U–test: P < 0.0001). All four individuals for which an autopsy was possible appeared to have died from a collision.

Discussion The translocation of wild hares appears to give very interesting results. Two years after release, about

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Table 4. Comparison of the home range sizes between various areas (Mann–Whitney U–test). Tabla 4. Comparación de los tamaños de las áreas de deambulación entre distintas zonas (test U de Mann–Whitney).

Source–Arenthon

Source–Mieussy

Arenthon–Mieussy

Kernel 95

P = 0.004*

P = 0.059

P = 0.56

Kernel 50

P = 0.012*

P = 0.12

P = 0.4

Total duration

After one month Kernel 95

P = 0.4

P = 0.14

P = 0.45

Kernel 50

P = 0.19

P = 0.14

P = 0.88

After two months Kernel 95

P = 0.88

P = 0.37

P = 0.39

Kernel 50

P = 0.51

P = 0.93

P=1

one fourth of the tagged individuals can still be observed during spot–light censuses. This is a very good result considering that this method is known to underestimate the real abundance of hare populations (Zellweger–Fischer et al., 2011). This is especially true for Mieussy, the target area located in the mountain, as the road network is very poor and the extent of the area sampled is thus limited. For this latter area, results are particularly interesting as hares originating from a cereal crop dominated lowland landscape were translocated to a pasture dominated mountainous area, and this usually in early winter, just before the first snowfall. Furthermore, during the mentioned censuses, conducted two years after the last release, the majority of the hares present were not tagged, which indicates an existing natural reproduction of the released individuals and of their offspring (and indeed we regularly observed leverets). The probability of hares migrating in those areas is very low because of a high degree of fragmentation and due to the absence of any hare populations in neighbouring areas. The individuals without tags are therefore most likely to be born from released hares. The mean duration of tracking of hares fitted with transmitters also indicates good survival. Means in Arenthon and Mieussy are close to those obtained from the source population. Finally, the released hares settled after only 20 days in Arenthon (the lowland area) and 50 days in Mieussy (the area located in the mountain), indicating how adaptable wild animals can be. The longer time to stabilisation observed in Mieussy is likely linked to the difference in environmental conditions between the source area and this target area. Translocations are often considered to have a low rate of success, particularly because of the induced stress during capture, handling, captivity, transport, release, and acclimation in the release sites leading to a lowered survival (Letty et al., 2003; Pelorosso et

al., 2008; Dickens et al., 2010). We tried to reduce this stress to a maximum by reducing the length of each step. Between capture and release we had a mortality rate of 5.9%. Even if it is not possible to give a precise mortality rate for the period after release, the settlement period, our results (re–observation rate after two years, mean tracking duration) suggest a rather high survival rate. An additional concern with translocated animals is the dispersion out of the target area (Dickens et al., op. cit.; Ferretti et al., 2010). In this respect, we had again rather good outcomes with less than 15% moving out of the selected areas, and all remaining in directly adjacent areas. For a translocation programme to be considered successful, it is also important to measure the impact of captures on the source population. In our project, the stability of densities observed in the source area indicates that the disturbances due to capture events and the withdrawal of 170 individuals within three years did not threaten the local population. Populations of hares can be quite robust when living at high densities. The adaptability of wild hares is also shown by the evolution of home range use of the released individuals. Although Ferretti et al. (2010) observed significantly larger home ranges in translocated hares as compared to residents, this was not the case in our study. Actually, in our study the difference was significant when considering the total dataset, but not anymore when taking in account that hares need a period of acclimation after release that was determined to be between one and two months according to our results. The mentioned authors did not consider this period of acclimation. Stress might be a factor that impairs the stabilisation of released individuals as its effects could reduce their ability to remember the location of vital resources (Teixeira et al., 2007). Furthermore, hares that roam more extensively are more exposed to predation or road casualties and have higher energy

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expenditure (Ferretti et al., op. cit.). Thus, stabilisation of home ranges, with a size comparable to that observed in the source population, indicates that released individuals have adapted to their new environment and that the target areas are therefore suitable. Comparing our results to other studies conducted with similar effort and in similar landscapes, home ranges in Geneva and the target areas are slightly smaller (Marboutin & Aebisher, 1996) or of similar size (Reitz & Léonard, 1994; Kunst et al., 2001). The value of the results we obtained in our study is enhanced when compared with the poor results recorded with the commonly used release of captive bred individuals. The low performance of this method is striking despite the low number of captive–bred hares we tracked (N = 8). High mortality rates during the first days after release were also reported by other authors for the Brown hare (Marboutin et al., 1990; Angelici et al., 2000). The better success obtained by releasing wild hares rather than individuals bred in captivity has already been documented in other studies (Pépin & Cargnelutti, 1985). So as expected, translocating wild individuals appears to be much more sustainable. It is also more ethical. There are, however, some constraints that need to be dealt with, such as the availability and proximity of a potential source population, and the availability of enough manpower for the captures. Regarding sustainability, the question arises of how the populations will evolve and survive in the long term in the relatively isolated areas that were restocked in our project. In the absence of an extension of the released population and of exchanges with other populations allowing the introduction of 'new blood' to improve genetic diversity, sustainability remains questionable (Fulgione et al., 2009). Managers in the target areas have now to improve habitat suitability for hares and the connectivity to neighbouring areas. Acknowledgements We are most grateful to Gottlieb Dändliker and Franck Péray of the Direction Générale Nature et Paysage (Genève), Guillaume Coursat, Eric Coudurier, and Pascal Roche of the Fédération des Chasseures de Haute–Savoie, and Laurent Loze, from Mieussy, for their valuable help throughout this project. Many thanks also to two anonymous referees for their useful comments on the first draft of our manuscript. References Angelici, F. M., Riga, F., Boitani, L. & Luiselli, L., 2000. Fate of captive–reared brown hares Lepus europaeus released at a mountain site in central Italy. Wildlife Biology, 6: 173–178. Baldi, A. & Farago, S., 2007. Long–term changes of farmland game populations in a post–socialist country (Hungary). Agriculture, Ecosystems and Environment, 118: 307–311. Benmergui, M., Reitz, F. & Fiechter, A., 1990. Taux de reprise et dispersion de lièvres (Lepus europeaus)

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First estimation of Eurasian lynx (Lynx lynx) abundance and density using digital cameras and capture–recapture techniques in a German national park K. Weingarth, C. Heibl, F. Knauer, F. Zimmermann, L. Bufka & M. Heurich Weingarth, K., Heibl, C., Knauer, F., Zimmermann, F., Bufka, L. & Heurich, M., 2012. First estimation of Eurasian lynx (Lynx lynx) abundance and density using digital cameras and capture–recapture techniques in a German national park. Animal Biodiversity and Conservation, 35.2: 197–207. Abstract First estimation of Eurasian lynx (Lynx lynx) abundance and density using digital cameras and capture–recapture techniques in a German national park.— Eurasian lynx are individually identifiable by their unique coat markings, making them ideal candidates for capture–recapture (CMR) surveys. We evaluated the use of digital photography to estimate Eurasian lynx population abundance and density within the Bavarian Forest National Park. From November 2008 to January 2009 we placed 24 camera trap sites, each with two cameras facing each other (on well–used walking tracks). The units were placed based on a systematic grid of 2.7 km. We captured five independent and three juvenile lynx and calculated abundance estimates using Program Mark. We also compared density estimates based on the MMDM method (Mean Maximum Distance Moved) from telemetry data (½MMDMGPS) and from camera trapping data (½MMDMCAM). We estimated that in an effectively sampled area of 664 km2 the Eurasian lynx density was 0.9 individuals/100 km2 with ½MMDMCAM. The Eurasian lynx density calculated with ½MMDMGPS was 0.4 individuals/100 km2 in an effectively sampled area of 1,381 km2. Our results suggest that long–term photographic CMR sampling on a large scale may be a useful tool to monitor population trends of Eurasian lynx in accordance with the Fauna–Flora–Habitat Directive of the European Union. Key words: Lynx lynx, Camera trap, Capture–recapture, Abundance, Half MMDM, Actual MMDM, Density. Resumen Primera estima de la abundancia y de la densidad del lince euroasiático (Lynx lynx) utilizando cámaras digitales y técnicas de captura–recaptura en un parque nacional alemán.— Al lince euroasiático se le puede identificar individualmente mediante las marcas de su pelaje, que son únicas, lo que le convierte en un candidato ideal para los estudios de captura–recaptura (CMR). Hemos evaluado el uso de la fotografía digital para estimar la abundancia y la densidad de la población del lince euroasiático en el Parque Nacional Forestal Bávaro. Desde noviembre del 2008 a enero del 2009 establecimos 24 lugares de trampeo, cada uno de ellos provisto de dos cámaras encaradas entre sí, en lugares de paso frecuentados. Colocamos las unidades basándonos en una cuadrícula sistemática de 2,7 km. Capturamos cinco linces independientes y tres jóvenes, y calculamos las estimas de abundancia utilizando el programa Mark. También comparamos las estimas de densidad mediante el método MMDM (distancia media máxima recorrida) de datos telemétricos (½MMDMGPS) y de datos de las cámaras trampa (½MMDMCAM). Hallamos que en un área muestreada eficazmente de 664 km2 la densidad del lince euroasiático era de 0,9 individuos/100 km2 mediante ½MMDMCAM. La densidad del lince euroasiático calculada mediante el método ½MMDMGPS fue de 0,4 individuos/100 km2 en una zona muestreada eficazmente de 1.381 km2. Nuestros resultados sugieren que un muestreo fotográfico CMR a largo plazo y a gran escala puede ser una herramienta muy útil para monitorizar las tendencias poblacionales del lince euroasiático, según la Directiva de Hábitat, Flora y Fauna de la Unión Europea. Palabras clave: Lynx lynx, Cámara trampa, Captura–recaptura, Abundancia, Media MMDM, MMDM real, Densidad. Received: 6 II 12; Conditional acceptance: 23 IV 12; Final acceptance: 12 VI 12

ISSN: 1578–665X

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Kirsten Weingarth & Felix Knauer, Inst. for Forest Zoology, Fac. for Forest and Environmental Sciences, Univ. of Freiburg, Tennenbacherstr. 4, D–79106 Freiburg, Germany.– Kirsten Weingarth, Christoph Heibl & Marco Heurich, Bavarian Forest National Park, Dept. of Research and Documentation, Freyunger Str. 2, D–94481 Grafenau, Germany.– Fridolin Zimmermann, KORA–Carnivore ecology and wildlife management, KORA, Thunstrasse 31, CH–3074 Muri, Switzerland.– Felix Knauer, Res. Inst. of Wildlife Ecology, Univ. of Veterinary Medicine Vienna, Savoyenstrasse 1, A–1160 Vienna, Austria.– Luděk Bufka, Šumava National Park Administration, 1 maje 260, CZ–35801 Vimperk, Czech Republic. Corresponding author: Kirsten Weingarth. E–mail: KirstenWeingarth@gmx.de

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Introduction How can we count a cryptic camouflaged species, with home range sizes up to 700 km2, in a low range mountain area? The Eurasian lynx is a secretive and elusive species that is difficult to monitor, but to implement management plans, wildlife managers need to know the size of wildlife populations. To date, monitoring of Eurasian lynx in Germany has been limited to chance observations and occasional telemetry studies, but these methods are unsuitable to obtain accurate abundance and density estimates. The individual coat markings and the behaviour of the Eurasian lynx make it an ideal candidate for systematic monitoring using remote photography and statistical capture–recapture methods (Cooch & White, 2006). In recent years, the use of camera traps has been implemented to estimate abundances of individually recognisable species such as felids. e.g., with tigers (Karanth & Nichols, 1998), ocelots Leopardus pardalis (Trolle & Kéry, 2003), jaguars Panthera onca (Silver et al., 2004), Iberian lynx Lynx pardinus (Gil–Sánchez et al., 2011) and bobcats Lynx rufus (Larrucea et al., 2007). The challenge of camera trap monitoring is to maximize the number of target species captures by assuring that every individual has the chance to be detected. This means that every potential home range should include camera trapping sites. For species like the Eurasian lynx, which presumably occur in low densities, site selection is critical to obtain a sufficient number of pictures. Therefore, in addition to a suitable site it is crucial to find a reliable camera trap that can deliver high quality pictures that will allow individual recognition. The Eurasian lynx population of the Bavarian and Bohemian Forest was newly founded in the 1980s following lynx releases in the area that is now the Šumava National Park, Czech Republik (Bufka & Cerveny, 1996). Sources of information concerning the progress of the population mainly came from unconfirmed references (Wölfl et al., 2001). In 1996 the Czech National Park Šumava set up the first telemetry projects, and in 2000 German telemetry projects were launched to support this initiative and thirteen Eurasian lynx were collared (Heurich & Wölfl, 2002; Bufka & Cerveny, 1996). Radio–telemetry delivers high–quality data, but it is invasive and costly (Gil–Sánchez et al., 2011). It mainly captures movement and behaviour although other information can be obtained, such as, kill rates for carnivores. Although there has been evidence of reproduction in the study area, it was seldom possible to capture dispersal or life histories of any animals other than the collared animals. Information regarding Eurasian lynx numbers, required by the lynx monitoring plan of the state of Bavaria, was still lacking (StMUGV, 2008). Abundance and density estimates of Eurasian lynx are required as a key factor to understand life histories and demography for decision–making in conservation (e.g., Fauna–Flora– Habitat directive) and politics (Hetherington & Gorman, 2007; Andrén et al., 2006). Digital camera traps offer a non–invasive, less costly method to evaluate the status of the Eurasian lynx population. Camera traps could allow us to monitor lynx demography by following indi-

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vidual life histories and assessing survival, recruitment and even dispersal. With this objective, we set up the first camera trap monitoring in a German National Park to test whether it is possible to generate abundance and density estimates in the putative core area of the Eurasian lynx population in the Bavarian Forest. Study area The Bohemian Forest and the Inner Bavarian Forest form one of the largest connected woodlands in Central Europe: The Greater Bohemian Forest Ecosystem is the largest, strictly protected, contiguous forest expanse in Central Europe. Entire tracts of forest are the property of the Bavarian state or the Czech Republic. The region is characterized by a low density of human habitation compared to other parts of Europe. In the core areas, this density it is less than 30 inhabitants/km², with approximately 70 inhabitants/km2 at the margins. Vast parts of this expanse are protected areas, such as the German Bavarian Forest National Park (with 242 km2) and the Czech Šumava National Park (with 690 km2) (Heurich & Wölfl, 2002), both surrounded by landscape protected areas. We conducted research in the IUCN Category II Bavarian Forest National Park with more than 98% forest cover (Elling et al., 1987). This area is located in the centre of this complex, extending along the Czech border. Forestry had been the dominating form of land use until the National Park was founded in 1970. Altitudes range from 650 m to a maximum of 1,420 m. The climate of the Bavarian Forest National Park is characterized by Atlantic and continental influences. The total annual precipitation is between 1,200 and 1,800 mm depending on altitude. Up to 50% of this amount falls as snow and the snow heights in the highest parts can reach up to 3 m (Bässler et al., 2008). Annual mean air temperature varies from 3.8°C in the high montane zones to 5.8°C in the valley sites (Noack, 1979; Bässler, 2004). The lowest temperature during the camera trapping session was reached in January with –12.4°C. There was snow from 22th of November until 10th of April and the snow level was highest in February with 111 cm at 945 m above sea level (weather station Waldhäuser). The National Park is a popular tourist site in summer and winter. There are 215 km of bike routes, 351 km of hiking trails 75 km being official winter hiking trails —and 85 km of cross–country skiing routes in use. Material and methods Camera traps The technique of individual recognition is based on the unique coat pattern of every Eurasian lynx (Karanth & Nichols, 1998; Karanth, 1995; Thüler, 2002; Garrote et al., 2011; Gil–Sánchez et al., 2010; Gil–Sánchez et al., 2011; Larrucea et al., 2007). For the accurate comparisons of individuals high quality pictures of both sides of the flanks are needed, including the inner surfaces of the fore and hind legs (Silver et al.,

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Table 1. Names, sex and transmission dates for seven individuals of Eurasian lynx (Lynx lynx) radio–tracked in the study area between 2008 and 2012. The transmission of ‘Milan’ covered two camera trapping sessions; the other individuals were radio–tracked during one camera trapping session: S. Sex (M. Male; F. Female); D. Transmission duration (in days): O. Ongoing. Tabla 1. Nombre, sexo y datos de transmisión de siete individuos de lince euroasiático (Lynx lynx) a los que se hizo un radio–seguimiento en el área de estudio entre 2008 y 2012. La transmisión de ''Milan'' se solapó con dos sesiones de cámara trampa; los demás individuos estaban siendo seguidos durante una sola sesión de cámara trampa: S. Sexo (M. Macho; F. Hembra); D. Duración de la transmisión (en días): O. En curso.

Transmission

Individual S Milan

Off

M 12 XI 2008 13 II 2010

D 458

Matilda F

17 III 2010 01 III 2011

349

Kubicka F

17 III 2010

07 II 2011

327

Ctirad

15 I 2011

14 III 2012

424

Tessa

27 II 2011

10 III 2012

377

Matilda F

02 III 2011

Kika

22 III 2011

2004). An initial trial of six camera models identified a passive infrared–triggered camera trap with white flash as the best in regard to image quality for use in the field (Cuddeback Capture Green Bay, Wisconsin, USA – Weingarth et al., in press). Due to the white flash the exposure time is shortened, resulting in sharp and fixed images with a very fine image definition. Consequently, the coat patterns of the Eurasian lynx can be distinguished without deforming the spots (Laass, 1999). The fast trigger speed of 0.3 sec is essential for use on trails if the animal is to be pictured in the centre of the image. The cameras ran for 24 h during the session and the delay between two pictures was set at a minimum of 30 sec. Telemetry The Eurasian lynx project of the Bavarian Forest National Park and Šumava National Park started in 2005, with a focus on the predator–prey relationships of Eurasian lynx and roe deer, and Eurasian lynx population trends in a low mountain area. Eurasian lynx are captured in wooden, two–door boxtraps (2.5 × 1 × 1 m), which are set up along forest roads and hiking paths used by the animals as trails. The traps are monitored continually with an electric

transmitter that sends a message by SMS. Sedation is achieved by shooting through a closable opening in the trap with a blowpipe and Hellabrunner mixture (Heurich, 2011). The Eurasian lynx were equipped with GPS–GSM collars (Vectronic Aerospace, Berlin, Germany). The collars were programmed for two daily fixings at 12:00 am and 12:00 p.m. Table 1 shows the dataset of Eurasian lynx that were have been equipped with collars during the 60–day period of the camera trapping session (26.11–24.01) over the years. We used telemetry data from previous years of the camera trapping study, to have a sufficient number of animals (N = 7) for the analysis. This was possible, because we assumed a constant Eurasian lynx density from snow tracking data. Study design Systematic distribution The distribution of the traps was designed to ensure that every individual in the study area had the chance of being detected (Karanth & Nichols, 1998). Therefore, a camera trapping site was set up in every second grid cell with an edge length of 2.7 × 2.7 km for a systematic distribution according to Laass (1999). This resulted in four to five camera trapping sites within an average female home range (Karanth & Nichols, 2002). Two opposing cameras were installed parallel to each other and 70 cm above the ground (withers of Eurasian lynx) to record both flanks (Silver et al., 2004). We installed 48 cameras, on 24 sites for the first intensive camera trapping session in the Bavarian Forest National Park in November 2008 (fig. 1). Each opposing pair of cameras was installed at a distance of 4.5 to 10 m and turned slightly away from each other to avoid interaction of the flashes and overexposure of the image. The camera traps were installed in wooden covers as a shelter against physical damage. The height of the camera was adjusted to the snow height by shifting it up and down a wooden pole. The minimum convex polygon (MCP; fig. 1) of all camera trapping sites formed a study area of 275 km2. Site selection and control routine For the site selection we displayed the telemetry data of two former collared Eurasian lynx, added the systematic snow tracking data since 1997, accidental lynx observations (tracks, kills, vocalisations, visual observations) and lynx prey sites since 2005 in a geographic information system (ArcGIS 9.3). Due to analysis of prey selection in the National Park Bavarian Forest, we assume that roe deer Capreolus capreolus is the most important prey species in the area as it is elsewhere in Central Europe (Okarma et al., 1997; Molinari–Jobin et al., 2007). Therefore, telemetry data of 64 roe deer collared in the study area were also included. Additionally, local and international experts selected trap locations based on their experience and topographical aspects. For example, rocky areas are preferred by Eurasian lynx for day resting sites and

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Šumava National Park MCP Grid (2.7 x 2.7 km) German–Czech border BFNP SNP 0

3.5

10.5 km

Bavarian Forest National Park

Fig. 1. Map of the Bavarian Forest National Park (BFNP) and Šumava National Park (SNP) showing the grid (2.7 × 2.7 km) used to position the 24 camera trapping sites ( ). The study area was defined as the minimum convex polygon (MCP) of the camera trapping sites. Fig. 1. Mapa del Parque Nacional Forestal Bávaro y Parque Nacional Šumava, mostrando la cuadrícula (2,7 x 2,7 km) utilizada para situar el emplazamiento de las 24 cámaras trampa ( ). El área de estudio se definió como el polígono convexo mínimo (MCP) de los emplazamientos de las cámaras.

chances are high that lynx use trails along ridges. To determine the exact site we relied on expert advice and locations that had a high density of data. Practical considerations, however, limited site selection. Sites above 1,200 m were excluded because of costly maintenance (low infrastructure, high snow levels) during the snow season. This is justified by the telemetry data of Eurasian lynx and roe deer in the study area, which shows low usage of the high elevations in winter. For the site selection, topography and vegetation structures were also taken into consideration as possible Eurasian lynx marking spots, tree cover and potential daily resting sites (Matjuschkin, 1978). Locations that lend themselves as easy passes, such as tree trunks over rivers or ridges leading to marking spots (Karanth & Nichols, 1998), can be of advantage. We controlled the camera trapping sites once a week so as to solve any technical failures, to adapt the camera positions to changing snow conditions, to check the alkaline batteries (variation in temperatures between +10°C in the sun until –15°C at night), and to assure no loss of pictures. A trap night was defined as effective if at least one camera at the site was able to produce images. The term 'potential trap night' means that the cameras were theoretically able to produce photos. If potential trap nights are not effective, influences such as snow in front of the lenses, defective flashes or low batteries prevented both cameras to detect objects.

Time of operation For this first camera trapping monitoring, we chose a session length of 60 days, (Karanth & Nichols, 1998, 2000; Guil et al., 2010). The length of one trapping occasion was set to five days (Zimmermann et al., 2008), i.e., several captures of the same individual at one particular camera trap site during five days are counted as a single capture event. The monitoring was carried out during the winter season because of positive experiences in Switzerland with less human disturbance in winter time. Additionally, between November and March, male Eurasian lynx have to cover long distances to find females and induce ovulation with their visits and defend their territories against other males during pre–mating season (Breitenmoser et al., 2006; Zimmermann et al., 2004). Due to snow tracking (Heurich et al., 2003) we know that Eurasian lynx in the Bavarian Forest National Park often frequent established routes, probably because it is the easiest way to move from A to B (Zimmermann et al., 2004). We assumed that touristic used winter hiking trails and snow hiking trails would offer an adequate chance to detect Eurasian lynx on the trail. Visual identification Like other felids (Trolle & Kéry, 2003, Karanth & Nichols, 1998), Eurasian lynx can be identified by their individual fur patterns, which they maintain their whole lifetime

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Fig. 2. Coat pattern of Eurasian lynx (Lynx lynx) used in the recognition of individual animals: A. A male lynx during sedation; B. The same individual on a camera trap image. For visual identification we compared three patches of the coat pattern (red ovals) to be discernible and congruent (Laass, 1999). Fig. 2. Patrones de manchas del pelaje de un lince euroasiático (Lynx lynx) utilizados para el reconocimiento de los animales individuales: A. Un lince macho sedado; B. El mismo individuo en una imagen de la cámara trampa. Para la identificación visual comparamos tres zonas del dibujo del pelaje (óvalos rojos) para que el reconocimiento fuera discernible y congruente (Laass, 1999).

(Guil et al., 2010). Therefore, we compared three different regions of the body, particularly the flanks or the inner legs (fig. 2; Laass, 1999). Sexual determination is only possible if a female is photographed with kittens or by detection of the nether regions (Guil et al., 2010). Age of the individuals cannot usually be determined exactly. Therefore, we defined three categories for the status of each photographed individual: The first category was 'independent' Eurasian lynx; this included adult and resident lynx identified through capture for GPS–collaring, animals that were documented for at least two years in the area, and lynx with cubs on camera trapping pictures. The 'independent' category would also include animals which were definitely over one–year old (subadults), when evidence was present in forms of camera trapping pictures taken in juvenile status one year ago (i.e., year of birth is known; Rexstad & Burnham, 1991). The second category describes 'juveniles', which are still dependent on the mother.

We defined the first 'lynx–year' from May 1 to April 30 of the following year when individuals start to disperse (Zimmermann et al., 2005). The third category, Eurasian lynx of 'unknown status', encompasses all remaining individuals without proof of independence or residency. Statistical analysis We tested the assumption of a closed population using CloseTest (Stanley & Burnham, 2004). A closed population means that there is no emigration, immigration, natality or mortality of individuals during the session duration. The captures and recaptures of Eurasian lynx were described by a binary matrix. Following Karanth & Nichols (1998), we defined five days to be one trapping occasion. We used closed population models in Mark (White & Burnham, 1999) for the abundance estimates. The model selection in Program Mark proposes the most appropriate model for the data.

Table 2. Results of the model selection in Mark. The model indices mean constant capture probability (o); capture probabilities vary by individual (h); capture probabilities vary by behavioral response to capture (b) and capture probabilities vary with time (t). Selected model has the maximum value. Tabla 2. Resultados de la selección de modelo en Mark. Los subíndices del modelo significan: probabilidad de captura constante (o); las probabilidades de captura varían según el individuo (h); las probabilidades de captura varían según la respuesta conductual a la captura (b); y las probabilidades de captura varían con el tiempo (t). El modelo seleccionado es el de valor máximo. Model Criterion

Mbh

Mth

Mtb

Mtbh

0.95

1.00

0.71

0.79

0.00

0.37

0.75

0.69

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To estimate density we applied mean maximum distance moved (MMDM) measures as a buffer around the study area in order to obtain the effective sampled area. Originally, MMDM was based on camera trap data (hereafter MMDMCAM) which is dependent on the camera trap design. MMDMCAM cannot be greater than the largest distance between two camera trapping sites. If the individual movement pattern of the species in concern includes larger distances, this might lead to overestimation of density. MMDM based on telemetry data (called 'actual' MMDM by Soisalo & Cavalcanti, 2006; hereafter ½MMDMGPS) might be a better option (Karanth, 1995; Soisalo & Cavalcanti, 2006), because the realisation of GPS locations is not confined to the study area. Here, we compare two measures, the ½MMDMCAM, which has often been used for rare felids (Karanth et al., 2002; Karanth et al., 2004), and the ½MMDMGPS. Results

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Table 3. The maximum distances moved (MDM, in km) by collared animals from 2008 to 2012. Tabla 3. Máximas distancias recorridas (MDM, en km) por los animales provistos de collar de 2008 a 2012. Lynx individual

Season

MDM

Milan

2008/2009

37.36

Milan

2009/2010

33.95

Kubicka

2010/2011

11.91

Matilda

2010/2011

12.95

Kika

2011/2012

23.73

Matilda

2011/2012

13.14

Ctirad

2011/2012

18.19

Tessa

2011/2012

10.60

Capture success and camera efficiency We found 1,414 out of 1,440 potential trap nights on 24 sites with 48 cameras over 60 days to be effective (98.2%). Two cameras were stolen but they were immediately replaced during the camera trapping session. We obtained 26 images of Eurasian lynx corresponding to a trapping rate of 1.8 lynx images/100 trap nights. During the camera trapping session we took photos of five independent individuals (two males and three females) and three juvenile individuals (sex unknown). Ten out of 24 sites were frequented by Eurasian lynx (41.6%). The family relations between the detected Eurasian lynx kittens and their mothers were obvious due to very small time intervals (< 5 min) between the detections on sites within the mothers´ home ranges. Following the same logic, subsequent images of juveniles without their mother were counted as a recapture of their mother (Zimmermann et al., 2004). We had eleven captures in total and four independent Eurasian lynx were recaptured, a female with a maximum of three recaptures. The amount of failed photos was < 5%. Abundance estimation The Close Test resulted in significance level of p = 0.05764, which means demographic closure is assured during the session. The minimal count within 60 days was five independent individuals which were the basis of our calculation. The model selection of program Mark selected the Mh model as the most appropriate (table 2). The mean value of 12 trapping occasions was six (CI: 6–15). The average capture probability is p = 0.1528 (Otis et al., 1978), with standard error 1.7440. Density estimations Four independent Eurasian lynx frequented at least two camera trapping sites. The maximum distances

moved ranged from 3.67 km (female) to 11.38 km (male). The ½MMDMCAM of 4.28 km (N = 4) resulted in an area effectively sampled of 664 km2 (MCP study area: 275 km2). Based on our abundance estimate of six independent individuals, this corresponds to a density of 0.9 independent individuals per 100 km2. From the GPS data of seven Eurasian lynx radio–tracked within the period of the camera trapping session (60 days) in the study area (table 1), we obtained eight maximum distances moved (table 3; the transmission duration of 'Milan' covered two camera trapping sessions) and a ½MMDMGPS of 10.12 km for the buffer radius (fig. 3). The effective sampled area is 1,381 km2, giving an estimate of 0.4 lynx individuals/100 km2. Discussion Camera model and study design The Cuddeback Capture™ worked reliably during the whole winter session, with minimum temperatures of –12°C. The excellent picture quality with white flash enabled us to identify every individual on the images. The amount of failed images was very low ( > 5%) in relation to the large amount of high quality images and compared to earlier felid projects that had percentages from 32% to 75% (Jackson et al., 2005). Effective trap–nights More than 98% of potential trap nights during the session of 60 days were effective. This value lies in the upper range of comparable camera trapping effectivity of 84.2% (Jura North, winter of 2006/2007) and 97.9% in Switzerland (Northwestern Swiss Alps, winter 2009/2010; Zimmermann et al., 2011). The combination

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MCP ½MMDMCAM ½MMDMGPS BFNP SNP State forest 0

3.5

10.5 km

Fig. 3. Map showing the study area (black solid line) and two estimates for the effective study area obtained with a buffer radius of ½MMDMCAM (black dashed line) and ½MMDMGPS (grey solid line). Fig. 3. Mapa que muestra el area de estudio (línea continua negra) y dos estimas del área de estudio efectiva, obtenidos con un radio–buffer de ½MMDMCAM (línea discontínua negra) y ½MMDMGPS (línea continua gris).

of high quality images and low camera failure technically minimizes the risk of missing individuals. Based on the grid of 2.7 × 2.7 km, we covered the whole area systematically, so we can assume that every individual present in the study area had the chance of being detected. This is also suggested by the finding that all individuals equipped with a radio–tracking collar that were present in the area in 2008/2009 were detected. Camera traps on 41.6% of the 24 sites successfully detected individuals of Eurasian lynx, compared to 24% in the Jura (winter of 2007/2008; Zimmermann et al., 2007) and 65% in the Northwestern Swiss Alps (winter of 2007/2008; Zimmermann et al., 2008) using the same study design. These values reflect the fact that the mountainous topography of the Bavarian Forest National Park and the Jura offer less forced trails compared to an alpine topography in the Swiss Alps with its larger and steeper slopes. Recognition of age on camera trapping pictures In contrast to Guil et al. (2010), who studied Iberian lynx (Lynx pardinus), we are not convinced that the age of Eurasian lynx can be distinguished visually due to the body size, beard and brush size, or facial characteristics. We think this depends heavily on the season, as for example, a cub photographed in November can still be distinguished due to smaller body size. But this is difficult to achieve with a single individual taken in March. A former year kittens' body size at that time of the year is almost as big as a full–grown individual. In consequence we de-

fined three categories which are strictly evidence–based. Due to continued camera trapping we will also be able to recognize individuals on a more detailed basis (e.g. year of birth or sex) in consecutive years. Abundance estimate A camera trapping session during the pre–mating season of Eurasian lynx, when especially males show enhanced activity and visits of individuals from outside the study area are most likely (Breitenmoser & Breitenmoser– Würsten, 2008), cautions against the assumption of a demographically closed population. Nevertheless, the Close Test (Stanley & Burnham, 2004) did not reject the assumption of population closure within 60 days from November to January. The rapid detection of all individuals within 25 days (corresponding to five trapping occasions; fig. 4) and the subsequent recapture of all individuals also suggest that we detected only regularly moving individuals. The software package Mark selected the Mh as the most appropriate model. This is a common finding in felids, which present large heterogeneity of individual capture probabilities (Kelly & Holub, 2008) due to their individual heterogeneity in capture probability. Future studies should determine the optimal length a session should be for the Eurasian lynx and which period of the year is most suitable for the camera trapping regarding the closure assumption, man power effort, and trap night efficiency. Whether the amount of Eurasian lynx captures during the late spring, summer and autumn season is sufficient for valuable estimates

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Density estimations Density estimation needs to take into account that individual home ranges might include areas outside the study area. The ½MMDMCAM method is widely used to estimate density for felids (Karanth & Nichols, 1998). The density estimate with the ½MMDMCAM resulted in 0.9 individuals/100 km2, corresponding to a density estimate from the Central Swiss Alps of 0.85 independent individuals/100 km2 (Zimmermann et al., 2004). As expected, our density estimate based on ½MMDMGPS (0.4 individuals/100 km2) was lower than that based on ½MMDMCAM, suggesting that the maximum distances moved by Eurasian lynx can be greater than the array of camera trapping sites, especially considering the elongate shape of the study area (fig. 1). These results are in congruence with those of Soisalo & Cavalcanti (2006) that deriving ½MMDMGPS from radio–tracking data leads to less biased densities. Eurasian lynx population sizes are influenced by various factors; Hetherington & Gorman (2007) emphasized the strong relationship between Eurasian lynx density and ungulate biomass. Based on hunting statistics we assume a low roe deer density in the Bavarian Forest National Park, and consider that this would not be able to sustain higher long–term densities of Eurasian lynx. In Białowieza Primeval Forest (Poland and Belarus) high prey densities result in higher Eurasian lynx densities with 3 independent individuals/100 km2 (Jedrzejewski et al., 1996). Due to the elongated shape of the study area and the low sample size (N = 4), the ½MMDMCAM is a less accurate measure than the ½MMDMGPS (based on N = 8), suggesting that a future enlargement of the study should aim at creating a more compact shape. Then, with increasing number of recaptures at more than one camera trap site, the density estimates become more robust. Successful camera trapping studies rely on well– trained and experienced staff (Sharma & Jhala, 2010) but, compared to radio–tracking studies, they are more cost–efficient and non–invasive (Gil–Sánchez et al., 2011). While the main goal of telemetry studies is to analyze the spatial and temporal behavior of the target species, the priority of systematic camera trapping is to estimate the abundance and density of the population.

12 10 Lynx captures

is questionable. The detection of the five independent individuals within the first five trapping occasions (fig. 4) and the additional finding that we detected all collared animals present in the study area favours our assumption that we detected most of the individuals present in the study area. On the other hand, the abundance estimate of six individuals within the area seemed to be close to reality, taking unconfirmed sightings and expert–confirmed prey sites into consideration. Likewise, the telemetry data also suggest free space for exactly one more Eurasian lynx home range within the study area. However, the minimum count of five independent Eurasian lynx as the basis for the abundance estimate, the large confidence interval of six to 15 and the low number of recaptures, led us to the conclusion that the study area needs to be enlarged.

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8 6 4 2 0

1 2 3 4 5 6 7 8 9 10 11 12 Trapping occasions Cumulated captures

Total caught

Fig. 4. Capture history of the independent Eurasian lynx. Juveniles were counted as recapture of their respective mother (Zimmermann et al., 2004). All individuals were detected within the first five trapping occasions. Fig. 4. Historial de capturas de linces euroasiáticos independientes. Los juveniles se contabilizaron como recapturas de sus respectivas madres (Zimmermann et al., 2004). Todos los individuos se detectaron durante los cinco primeros trampeos.

Comparing different methods used to calculate carnivore densities, Balme et al. (2009) found that camera trapping produces accurate but less precise estimates than telemetry data. Here we have shown that the two techniques function best when used to complement each other: The mark–recapture design relies on camera trapping, but additional information, e.g., the calculation of ½MMDMGPS comes from telemetry data. The Eurasian lynx is listed in the Habitats Directive of the European Union in Annex II IV, which requires surveillance of the conservation status of this species by the authorities. Our results suggest camera trapping as an adequate monitoring tool for this purpose and we intend to implement long–term camera trap monitoring, as drafted in the Eurasian lynx management plan of Bavaria/Germany (StMUGV, 2008). If used properly, 'camera trap surveys represent the best balance of rigor and cost–effectiveness for estimating abundance and density of cryptic carnivore species that can be identified individually' (Balme et al., 2009). Acknowledgements We want to thank Martin Gahbauer for his extraordinary support during site selection. We also want to thank

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Horst Burghart, Martin Horn and Lothar Ertl for their assistance during collaring and telemetry. The team of Bavarian Forest National Park was a great help with their expert advice in the material construction, provision of control teams of National Park rangers, and advice during site selection and logistics. Financial support was provided by the EU–programme Interreg IV (Ziel 3) and the Bavarian Forest National Park administration. References Andrén, H., Linnell, J. D. C., Liberg, O., Andersen, R., Danell, A., Karlsson, J., Odden, J., Moa, P. F., Ahlqvist, P., Kvam, T., Franzén, R. & Segerström, P., 2006. Survival rates and causes of mortality in Eurasian lynx (Lynx lynx) in multi–use landscapes. Biological Conservation, 131: 23–32. Balme, G. A., Hunter, L. T. B. & Slotow, R., 2009. Evaluating Methods for Counting Cryptic Carnivores. Journal of Wildlife Management, 73: 433–441. Bässler, C., 2004. Klimawandel–Trend der Lufttemperatur im Inneren Bayerischen Wald (Böhmerwald). Silva Gabreta, 14: 1–18. Bässler, C., Förster, B. & Müller, C. M. A. J., 2008. The BIOKLIM Project: Biodiversity Research between Climate Change and Wilding in a temperate montane forest–The conceptual framework. Waldökologie Online. Breitenmoser, U. & Breitenmoser–Würsten, C., 2008. Der Luchs–Ein Großraubtier in der Kulturlandschaft., Wohlen/Bern, Salm Verlag. Breitenmoser, U., Breitenmoser–Würsten, C., Arx, M. V., Zimmermann, F., Ryser, A., Angst, C., Molinari– Jobin, A., Molinari, P., Linnell, J., Siegenthaler, A. & Weber, J.–M., 2006. KORA Bericht, 33: Guidelines for the Monitoring of Lynx. Bufka, L. & Cerveny, J., 1996. The lynx (Lynx lynx L.) in the Sumava region, southwest Bohemia. Journal of Wildlife Research, 1: 167–170. Cooch, E. & White, G., 2006. Program MARK: a gentle introduction. Colorado State Univ., Colorado. Elling, W., Bauer, E. & Klemm, G. K., 1987. Klima und Böden. Wissenschaftliche Reihe, Nationalparkverwaltung Bayerischer Wald. Garrote, G., Pérez de Ayala, R., Pereira, P., Robles, F., Guzmán, N., García, F. J., Iglesias, M. C., Hervás, J., Fajardo, I. & Simón, M., 2011. Estimation of the Iberian lynx (Lynx pardinus) population in the Doñana area, SW Spain, using capture–recapture analysis of camera trapping data. European Journal of Wildlife Research, 57: 355–362. Gil–Sánchez, J. M., Moral, M., Bueno, J., Rodríguez– Siles, J., Lillo, S., Pérez, J., Martín, J. M., Valenzuela, G., Garrote, G. & Torralba, B., 2011. The use of camera trapping for estimating Iberian lynx (Lynx pardinus) homeranges. European Journal of Wildlife Research, 57: 1203–1211. Gil–Sánchez, J. M., Simón, M. A., Cadenas, R., Bueno, J., Moral, M. & Rodríguez–Siles, J., 2010. Current status of the Iberian lynx (Lynx pardinus) in eastern Sierra Morena, southern Spain. Wildlife

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München. Thüler, K., 2002. Spatial and temporal distribution of coat patterns of Eurasian lynx (Lynx lynx) in two re–introduced populations in Switzerland. KORA– Bericht. Muri, KORA. Troller, M. & Kéry, M., 2003. Estimation of ocelot density in the Pantanal using capture–recapture analysis of camera trapping data. Journal of Mammalogy, 84: 607–614. Weingarth, K., Zimmermann, F., Knauer, F. & Heurich, M., in press. Evaluation of six digital camera models for the use in capture–recapture sampling of Eurasian Lynx (Lynx lynx). Forest Ecology, Landscape Research and Nature Protection. White, G. & Burnham, K., 1999. Program MARK: survival estimation from populations of marked animals. Bird study, 46: S120–139. Wölfl, M., Bufka, L., Červený, J., Koubek, P., Heurich, M., Habel, H., Hubert, T. & Poost, W., 2001. Distribution and status of lynx in the border region between Czech Republic, Germany and Austria. Acta Theriologica, 46: 181–194. Zimmermann, F., Fattebert, J., Breitenmoser–Würsten, C. & Breitenmoser, U., 2007. Abundanz und Dichte der Luchse Fang–Wiederfang–Schätzung mittels Fotofallen im nördlichen Schweizer Jura. KORA–Bericht. Zimmermann, F., Fattebert, J, Caviezel, S., Breitenmoser–Würsten, C. & Breitenmoser, U., 2008. Abundanz und Dichte des Luchses in den Nordwestalpen Fang–Wiederfang–Schätzung mittels Fotofallen im K–VI. KORA–Bericht. Zimmermann, F., Molinari–Jobin, A., Capt, S., Ryser, A., Angst, C., Von Wattenwyl, K., Burri, A., Breitenmoser–Würsten, C. & Breitenmoser, U., 2004. Monitoring Luchs Schweiz 2003. KORA–Bericht. Muri, KORA. Zimmermann, F., Molinari–Jobin, A., Weber, J.–M., Capt, S., Ryser, A., Angst, C., Breitenmoser–Würsten, C. & Breitenmoser, U., 2005. Monitoring der Raubtiere in der Schweiz 2004. KORA–Bericht. Muri, KORA. Zimmermann, F., Werhahn, G., Hofer, L., Poole, S., Ryser, A., Breitenmoser–Würsten, C. & Breitenmoser, U., 2011. Abundanz und Dichte des Luchse in der Zentralschweiz West: Fang– Wiederfang–Schätzung mittels Fotofallen im K–III im Winter 2010/11. KORA–Bericht.

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Reproduction of wild boar in a cropland and coastal wetland area: implications for management C. Rosell, F. Navàs & S. Romero

Rosell, C., Navàs, F. & Romero, S., 2012. Reproduction of wild boar in a cropland and coastal wetland area: implications for management. Animal Biodiversity and Conservation, 35.2: 209–217. Abstract Reproduction of wild boar in a cropland and coastal wetland area: implications for management.— The reproductive parameters of a wild boar population located in a coastal landscape with a mosaic of cropland and wetland habitats were analysed and compared with those observed in wild boar populations living in other habitats. A total of 296 reproductive tracts of females captured year round at the Aiguamolls de l'Empordà Natural Park were collected and analysed from 2000 to 2010. The foetuses were counted, sexed and aged and the mating and birth periods were determined. The weight and age of each female were also recorded. In accordance with the pattern observed in most European populations, a marked main mating season from October to January was observed. Within this season, there was a peak during November and December, in which 64% of the conception dates were recorded. The proportion of breeding females, ovulation rate and litter size increased with the weight of the reproductive females. A mean litter size of 5.01 ± 1.33 (range from two to eight) foetuses was recorded. This value is the highest known litter size recorded in wild Iberian populations and is similar to values observed in central Europe. Furthermore, it is not in accordance with the pattern reported for other European populations in which a positive correlation between litter size and latitude was observed. The most likely explanation for the high reproductive output in the study area is the availability of food year round, and especially the high consumption of crops such as maize and sunflower. Our results suggest that colonisation of cropland and wetland areas is contributing to the rise in the wild boar population density. Control strategies should consider not only reducing numbers of adult females but also applying measures to reduce food resources available to wild boar. Key words: Wild boar, Sus scrofa, Reproduction, Litter size, Management. Resumen Reproducción del jabalí en hábitats de cultivos y humedales costeros: implicaciones sobre la gestión.— Los parámetros reproductivos de una población de jabalí localizada en un paisaje costero, con un mosaico de cultivos y humedales, fueron analizados y comparados con los observados en poblaciones de jabalí que colonizan otros hábitats. Se analizaron un total de 296 tractos reproductivos procedentes de hembras capturadas a lo largo de todo el ciclo anual en el Parque Natural de Els Aiguamolls de l'Empordà entre los años 2000 y 2010. Se contaron los fetos que presentaban las hembras gestantes y se determinó su sexo y edad, así como los períodos de copula y parto. También se registró el peso y edad de cada hembra. De acuerdo con el patrón observado en la mayor parte de poblaciones europeas de la especie, se observó un período de celo principal entre octubre y enero, con un máximo durante noviembre y diciembre, meses que concentraron el 64% de las cópulas. La proporción de hembras gestantes, la tasa de ovulación y el tamaño de camada aumentan con el peso de la hembra. El tamaño medio de camada registrado fue de 5,01 ± 1,33 (rango de dos a ocho) fetos. Este valor es el más elevado registrado en poblaciones salvajes Ibéricas y es parecido al observado en algunas poblaciones del centro de Europa. Además, no se corresponde con el patrón descrito para las poblaciones de jabalí en Europa según el cual se aprecia una correlación positiva entre el aumento del tamaño de la camada y la latitud. La explicación más probable para la alta productividad de la población en la zona de estudio es la gran disponibilidad de alimento a lo largo de todo el año y, especialmente, el elevado consumo de plantas cultivadas, particularmente maíz y girasol. Estos resultados sugieren que la colonización de zonas agrícolas ISSN: 1578–665X

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y humedales contribuye al aumento de densidad de población del jabalí y las estrategias de control deberían considerar tanto la reducción del número de hembras adultas como la aplicación de medidas para reducir la disponibilidad de recursos tróficos accesibles para el jabalí. Palabras clave: Jabalí, Sus scrofa, Reproducción, Tamaño de camada, Gestión. Received: 9 III 12; Conditional acceptance: 17 V 12; Final acceptance: 10 VII 12 Carme Rosell & Ferran Navàs, MINUARTIA, P/Domènech 3, 08470 Sant Celoni, Barcelona (Spain).– Carme Rosell, Dept. Biologia Animal (Vertebrats), Fac. Biologia, Univ. de Barcelona, Barcelona (Spain).– Sergi Romero, Dept d'Agricultura, Ramaderia, Pesca, Alimentació i Medi Natural, Parc Natural dels Aiguamolls de l'Empordà, ctra. Sant Pere Pescador km 13,6, 17486 Castelló d’Empúries, Girona (Spain). Corresponding author: Carme Rosell: E–mail: crosell@minuartia.com

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Introduction The wild boar (Sus scrofa L., 1758) has become a species of strategic social and economic interest. European populations have increased greatly in recent decades (e.g. Genov, 1981; Erkinaro et al., 1982; Saéz–Royuela & Tellería, 1986; Goulding et al., 1998; Melis et al., 2006; Apollonio et al., 2008). The species affects biodiversity and protected areas and has been highlighted as a vertebrate that can modify natural plant communities (e.g. Engeman et al., 2007; Muñoz & Bonal, 2007; Webber et al., 2010; Bueno et al., 2010). It also damages crops (e.g. Mackin, 1970; Goulding et al., 1998; Schley & Roper, 2003), affects forest regeneration and habitat restoration (Mayer et al., 2000; Gómez & Hódar, 2008), generates conflicts as a result of its presence in urban areas (Cahill & Llimona, 2004; Jansen et al., 2007), and causes traffic accidents due to collisions (Groot Bruinderink & Hazebroek, 1996; Colino et al., 2012). Wild boar also have the potential to transmit disease to livestock (Gortázar et al., 2006; Santos et al., 2009). Control of wild boar populations is thus a major challenge for wildlife managers, and conflict mitigation measures are required (for a review see Massei et al., 2011). The species has a higher reproductive potential than other ungulates. They show an early onset of puberty, a relatively short gestation period and high mean litter size (Mauget, 1972). This is a key ecological feature that, together with its opportunistic, omnivorous diet (for a review see Schley & Roper, 2003) and its adaptability to a high variety of landscapes, has allowed wild boar to expand its distribution and numbers. Previous studies showed that food availability, determined by habitat features, weather conditions and human influence (the provision of supplementary food and crops), plays a role in wild boar body weight and reproductive parameters (e.g. Matschke, 1964; Briedermann, 1971; Pepin et al., 1987; Gaillard et al., 1993; Fernández–LLario & Mateos–Quesada, 2005; Treyer et al., 2012). Weather factors affect the abundance, quality and accessibility of food. Snowfall and low temperatures in winter periods have been found to affect reproductive parameters in North European populations (Oloff, 1951) and the effect of summer drought has been reported in Mediterranean areas (Massei et al., 1996; Fernández–Llario & Carranza, 2000). The influence of diseases on wild boar fertility and reproductive outcome was found to be of minor interest in some German populations (Gethöffer et al., 2007), but ovulation rate was negatively related to the seroprevalence of some protozoan parasites in Spanish populations (Ruiz–Fons et al., 2006). A strong correlation between wild boar litter size and latitude has been reported in European populations, with litter sizes increasing by an average of about 0.15 offspring per degree of latitude (Bywater et al., 2010). The lowest average litter sizes, around three foetuses per female, are found in southern Spain (Fernández–Llario & Carranza, 2000) and maximum litter sizes, above six foetuses per female, have been reported in Hungary and Germany (Oloff, 1951; Náhlik & Sandor, 2003; Gethöffer et al., 2007). In the Iberian

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populations, a range from 3.05 (Fernández–Llario & Carranza, 2000) to 4.5 foetuses per reproductive female (Herrero et al., 2008) has been reported. In this study, we analysed the reproductive parameters of a wild boar population living in a coastal wetland habitat surrounded by cropland. Food availability here is high year round. We assessed whether the productivity of this population is similar to that observed in other Iberian populations inhabiting forest and bush habitats and investigated the influence of female age and weight. Furthermore, we discuss the application of the results to control strategies and adaptive management of wild boar populations. Material and methods The study was carried out in the Aiguamolls de l'Empordà Natural Park (42° 13' 28.09'' N, 3° 05' 34.92'' E Catalonia, NE Spain). This is a Mediterranean coastal wetland area that has been declared an Important Bird Area (IBA) and is included in the Natura 2000 network. It has a total surface of 4,824 ha and includes four Integral Nature Reserves. The area is at sea level and is located between the Muga and Fluvià Rivers. The climate is Mediterranean with a total precipitation of 600 mm and dry periods during the summer and the winter. The mean annual temperature is 21.5°C max and 10.7°C min. The typical habitats of coastal marshland areas are found in the Natural Park, including beaches with sand dunes (5% of the total surface), salt marshes with vegetation dominated by glasswort Arcthrocnemum fruticosum, cordgrass Spartina versicolor and rush Juncus maritimus (35%), coastal lagoons with brackish/salt and fresh water (10%) ,and reed beds covered by reed Phragmites australis and lesser bulrush Typha angustifolia (35%). Meadows are separated by trees and small forests of tamarisks (Tamarix gallica and Tamarix africana), and are flooded in some periods (15%). The Natural Park is surrounded by irrigated agricultural land where the main crops are sunflower Helianthus annuus, maize Zea mays, barley Hordeum vulgare and, to a lesser extent, fruit trees such as apple Malus sp. and small areas of rice fields Oryza sativa. The area is home to a wide range of vertebrates with species of high conservation value. Two ungulate species are present in the Natural Park: fallow deer Dama dama, which was introduced into the area in 1993, and wild boar. Wild boar was only an occasional visitor before 1990, but the population has expanded and it is now one of the most abundant mammals in the area. No supplementary food is given to the population. Population control is required as a result of the high concentration of wild boar in the nature reserves. Since 1998, this has been carried out by rangers in the reserve areas, as hunting is forbidden. A total of 296 female wild boars captured at the natural reserves were collected from 2000 to 2010. All harvested females were weighed (total weight was recorded before evisceration) and their age was estimated according to tooth eruption and replacement pattern (Matschke, 1967; Saenz de Buruaga et

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al., 1991). The reproductive tracts (uteri and ovaries) were collected, preserved in formol 5% and examined in the laboratory. Ovarian activity was recorded and the ovulation rate was estimated as the mean of corpora lutea found in both ovaries. The foetuses in the uterus of pregnant females were counted, weighed, measured and sexed (when possible). The foetal age was determined using the formula

was compared with theoretical distribution 1:1 using the chi–square test. For all the tests, significance was assumed when p < 0.05. Analyses were performed using the STATISTICA® (StatSoft Inc., Tulsa, USA) package. Results Fifty–two percent of the females analysed were pregnant. The main wild boar rutting period in the study area took place from October to January (80% of the conception dates were recorded in this period) and reached a maximum during November and December (64%). Corresponding to these data, a peak in the birth period was observed during March and April. However, births were recorded year round (fig. 1). Wild boar females reached puberty at a minimum age of six months (0.69% of total breeding females) and a minimum weight of 30 kg. The ovulation rate was 5.18 ± 1.46 (mean ± SD; n = 136) and ranged from 1 to 8. Litter size was 5.01 ± 1.33 (mean ± SD; n = 88) and ranged from two to eight, with four to six being the most frequent number of foetuses per female (fig. 2). An 11% intrauterine loss was recorded. No distorted foetal sex ratio was observed (1.13:1; c2 = 0.51; df = 1; p = 0.475). Age and weight of females were positively correlated (y = 32.89 + 0.84x; RSpearman = 0.882; p < 0.05). A positive linear correlation was observed between litter size and female weight (y = 2.54 + 0.04 x; RSpearman= 0.41; p < 0.05). When females were grouped into the three weight classes most commonly applied in population management, the reproductive parameters increased according to weight class (table 1). Significant differ-

A = 22.5378 + 0.2893 L where A is the age of the foetuses (in days) and L is the average length of the foetuses found in each uterus (Vericad, 1983). Using the female’s date of death and the estimated age of the foetuses, the conception and birth (120 days after copulation, Mauget, 1972) dates of each litter were determined and grouped in monthly periods. The litter size was defined by the mean number of foetuses found in the uteri of the pregnant females. Intrauterine mortality was determined following Mauget (1972) and Abaigar (1992) where IUM =

Ovulation rate – Litter size x 100 Ovulation rate

Conceptions and births (%)

Annual productivity for each weight class was estimated according to Mauget (1982), considering the proportion of breeding females and the mean litter size. Non–parametric tests were used, as the assumptions of normality (evaluated using the Shapiro–Wilk W test) of the data were not satisfied. Linear correlation between variables was tested using the Spearman R analysis. Differences in breeding status, ovulation rate and litter sizes between weight classes were analysed using the Kruskal–Wallis test. The sex ratio

35 30 25 20 15 10 5 0

J J Month

Conceptions

Births

Fig. 1. Female wild boar reproductive phenology in Aiguamolls de l'Empordà Natural Park (n = 296; 2000–2010 period). Fig. 1. Fenología reproductiva de las hembras de jabalí en el Parque Natural de Aiguamolls de l'Empordà (n = 296; período 2000–2010).

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5.01 ± 1.33 SD (n = 88)

Frequency (%)

25 20 15 10 5 0

4 5 Litter size

Fig. 2. Distribution of litter size of the wild boar population in the Aiguamolls de l'Empordà Natural Park. Fig. 2. Distribución del tamaño de camada de la población de jabalíes del Parque Natural de Aiguamolls de l'Empordà.

ences were observed in the percentage of breeding females (c2 = 10.48, df = 2, p < 0.01), ovulation rate (H2,N=129 = 41.65, p < 0.001) and litter size (H2,N=83= 16.52, p < 0.001). The offspring per 100 females varied from 159.59 foetuses in females under 50 kg to 427.05 in females above 70 kg.

Discussion The reproductive phenology in the study area showed marked seasonal breeding, with the main rutting period in autumn and winter and births clustered mainly in spring. A few females gave birth in other months of the year. Although the study population is

Table 1. Reproductive parameters of wild boar females considering the weight class. Females weighing over 89 kg were not considered due to the low number of cases: * Foetuses/100 females. Tabla 1. Parámetros reproductivos de hembras de jabalí considerando la clase de peso. No se han incluido las hembras de peso superior a 89 kg debido al escaso número de casos: * Fetos/100 hembras.

30–49 kg Breeding females (%)

42.22 (n = 90)

50–69 kg 64.13 (n = 92)

70–89 kg 76.67 (n = 60)

Ovulation rate Mean ± SD Median range

3.94 ± 1.28 (n = 34)

5.21 ± 1.09 (n = 56)

6.13 ± 1.36 (n = 39)

1–6

2–8

Litter size Mean ± SD Median Range Productivity*

3.78 ± 1.05 (n = 14)

5.0 ± 1.02 (n = 39)

5.57 ± 1.50 (n = 30)

2–6

3–7

2–8

159.59

320.65

427.05

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Rosell et al.

Table 2. Litter size recorded (average, A) in different European wild boar populations. Tabla 2. Tamaño de camada registrado (media, A) en distintas poblaciones europeas de jabalí. Location

Reference

Location

Reference

Germany

6.91 Gethöffer et al., 2007

Hungary

6.7

Náhlik & Sandor, 2003

Catalonia

5.01 This paper

Germany

6.5

Oloff, 1951

Aragon

4.5

Austria

5.8

Martys, 1982

(Ebro River valley)

Germany

5.49 Ahrens, 1984 in

Iberian populations Herrero et al., 2008

Burgos

4.3

Bywater et al., 2010

Sáez–Royuela, 1987

Aragon

4.25 Herrero et al., 2008

Germany

5.3

Stubbe & Stubbe, 1977

(Pyrenees)

Germany

5.3

Briedermann, 1971

Portugal

4.17 Fonseca et al., 2004

Luxembourg

5.3

Cellina, 2008

Andalousie

4.1

France

5.2

Servanty et al., 2007

Castilla–

3.91

Germany

5.11

Gethöffer et al., 2007

Italy

4.95 Boitani et al., 1995

Extremadura 3.88 Garzón, 1991 Catalonia

Abaigar, 1992 Ruiz–Fons et al., 2005 la Mancha

Poland

4.83 Fruzinski, 1995

Switzerland

4.8

Moretti, 1995

Extremadura 3.75 Fernández–Llario &

Italy

4.7

Monaco et al., 2010

Mateos–Quesada,

France

4.62 Mauget, 1972

2005

3.78 Rosell, 1998

Italy

4.6

Cappai et al., 2008

Aragón

France

4.6

Aumaitre et al., 1984

(Pyrenees)

France

4.47 Aumaitre et al., 1982

Andalousie 3.05 Fernández–Llario &

France

4.44 Dardaillon, 1984

Italy

4.2

Switzerland

4.17 Neet, 1995

Italy

3.88 Massei et al., 1997

3.27 Vericad, 1983

Carranza, 2000

Focardi et al., 2008

located in a mosaic of croplands and marshlands, the seasonal pattern recorded corresponds to patterns observed in Iberian populations in other habitats (see Sáez–Royuela, 1987; Rosell et al., 2001 for a review; Fonseca et al., 2004; Fernández–Llario & Mateos–Quesada, 2005; Fonseca et al., 2011) and to those observed in other European populations (Briedermann, 1971; Aumaitre et al., 1984; Moretti, 1995; Durio et al., 1995; Boitani et al., 1995). The oestrus of female wild boar is greatly influenced by photoperiod and food resources, and the breeding seasonality corresponds to the availability of energy rich foods that provide optimal nutritional conditions for sows at the end of summer and in early autumn (Mauget, 1982; Fernández–Llario & Mateos–Quesada, 1998; Gethöffer et al., 2007; Servanty et al., 2009). The seasonal pattern of reproduction was also observed when energy supply is artificially increased by baiting (Treyer et al., 2012). In the study area, energy foods are provided mainly by crops ––maize and sun-

flower–– that account for 37% of the total volume of wild boar stomach contents (Giménez–Anaya et al., 2008). These cultivated plants are mainly consumed from July to October (67.3% of stomach contents in July–August and 64.3% in September–October). Wild boar living in this cropland area consume agricultural plants instead of mast that is the energy–rich food mainly consumed at the beginning of the rutting period in many locations (Schley & Roper, 2003). The mean litter size observed in the Aiguamolls de l’Empordà population is the highest recorded in the Iberian peninsula and is similar to that observed in Central European countries (table 1; see Bywater et al., 2010 for a review). Litters of above 5 young/ reproductive female have been reported in Austria, Germany, northern France and Luxemburg. In these Central European areas, wild boar found optimal environmental conditions and a high availability of natural food resources. Furthermore, supplementary food is often provided (Cellina, 2008; Servanty et al.,

Animal Biodiversity and Conservation 35.2 (2012)

2009). In Italy, Poland, Switzerland and France, lower productivity with less than 5 foetuses per female were reported (table 1). In the Iberian populations, the mean litter size reported reached the minimum in southern Spain under drought conditions (3.05 foetuses per female; Fernández-Llario & Carranza, 2000) and in the Pyrenean habitats (3.27; Vericad, 1983). Iberian populations located mainly in forests and shrublands show a mean litter size of around 4 foetuses/ female and the maximum litter size (4.5) was reported at the population located in the River Ebro valley in a crop landscape (Herrero et al., 2008; table 1). The most likely explanation for the high litter size observed in the study area may be related to the high availability of energy–rich cultivated foods (maize and sunflower) and the opportunity to obtain stable food sources throughout the year. These foods are provided mainly by crops during summer and autumn and by marshland food resources (underground parts of the wetland plants and animal matter) in winter and spring (Giménez–Anaya et al., 2008). However, we cannot rule out possible effects such us some degree of crossbreeding with domestic pigs that may have influenced the reproductive output. Slight phenotypic evidence of possible crossbreeding (small parts of the coat with lighter tones than usual in wild boar) was observed in a small proportion of individuals captured (around 5%; unpub.lished data). This issue has not been analysed in other European populations with available reproductive data, and further investigations on chromosome polymorphism among the different wild boar populations would clarify its effect on reproductive output, as other authors have previously suggested (Gethöffer et al., 2007). The high mean litter size observed in the study area does not fit with the pattern described by Bywater et al. (2010), whereby there is a strong correlation between litter size and latitude for wild boar in Europe. The use of litter sizes predicted on the basis of latitude has been suggested by these authors for management approaches based on population modelling as those proposed by Bieber & Ruff (2005). Nevertheless, according to our data, factors related to human influence affecting wild boar populations (such as high availability of food provided by crops, baiting or crossbreeding with domestic pig) must also be considered in the use of this litter size pattern for demographic modelling. The female body mass shows a high effect on the reproductive parameters in the study area where an increase was recorded in all the reproductive parameters studied (proportion of breeding females, ovulation rate and litter size) according to weight class. A positive correlation between reproductive output and female weight and age was also reported in many populations (e.g. Mauget, 1972; Aumaitre et al., 1982; Sáez–Royuela, 1987; Fonseca et al., 2011) and it was related to the fact that resources available to reproduction increase with female weight and better physical condition (Fernández–Llario & Mateos–Quesada, 1998; Massei et al., 1996). The results suggest that the high availability of food may contribute to the high reproductive output of the study population and to the increase in population density in this area. This observation has implications

215

regarding population control strategies, suggesting that the capture of larger females may be effective in reducing the potential for population growth. Some authors, however, have proposed reduction in the number of piglets as the target for wild boar population control (Sodeikat, et al., 2005). Based on population modelling, it has also been suggested that reducing juvenile survival would have a substantial effect under good environmental conditions, whereas reducing numbers of adult females would be more effective in years when conditions were poor (Bieber & Ruf, 2005). According to these data, management strategies for reducing wild boar density in protected nature reserves should combine different control techniques. Some, such as trapping, should have the objective of capturing juveniles, whilst others should focus on selective culling of adult females that have the highest reproduction output. Moreover, other measures to reduce reproductive potential should be applied, such as preventing crossbreeding with domestic pig, avoiding supplementary feeding, and applying crop protection measures to reduce the food resources available to wild boar. Acknowledgements This project was financed by the Direcció General de Medi Natural of the Generalitat de Catalunya. The authors wish to thank the ‘Cos d’Agents Rurals’ rangers and all the staff at the Aiguamolls de l'Empordà Natural Park for their help in data collection, and Josep Maria Espelta, CREAF, Autonomous University of Barcelona for his help in data analysis. Two anonymous reviewers helped to improve the manuscript. The first author is a member of the Vertebrate Biology Quality Research Group 2009SGR43 at the Department of Animal Biology, University of Barcelona, which is supported by the Catalan government. References Abaigar, T., 1992. Parametres de la reproduction chez le sanglier (Sus scrofa) dans le sud–est de la Peninsule Ibérique. Mammalia, 56(2): 245–250. Aumaitre, A., Morvan, C., Quere, J. P., Peiniau, J. & Valet, G., 1982. Productivité potentielle et reproduction hivernale chez la laie (Sus scrofa scrofa) en milieu sauvage. Journées Recherche. Porcine en France, 14: 109–124. Aumaitre, A., Quere, J. P. & Peiniau, J., 1984. Influence du mileu sur la reproduction hivernale et la prolificité de la laie. In: Symposium international sur le Sanglier. Les colloques de l’INRA, 22. Ed. INRA, Publ. Toulouse. Apollonio, M., Andersen, R. & Putman, R. (Eds.) 2008. Ungulate management in Europe in the XXI century. Cambridge Univ. Press, Cambridge. Bieber, C. & Ruf, T., 2005. Population dynamics in wild boar Sus scrofa: ecology, elasticity of growth rate and implications for the management of pulsed resource consumers. Journal of Applied Ecology, 42: 1203–1213.

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Human–wildlife interactions C. Rosell & F. Llimona

Rosell, C. & Llimona, F., 2012. Human–wildlife interactions. Animal Biodiversity and Conservation, 35.2: 219–220. The nature of wildlife management throughout the world is changing. The increase in the world’s human population has been accompanied by a rapid expansion of agricultural and urban areas and infrastructures, especially road and railway networks. Worldwide, wildlife habitats are being transformed and fragmented by human activities, and the behavior of several species has changed as a result of human activities. Some species have adapted easily to urban or peri–urban habitats and take advantage of the new resources available. These data provide the context for why human–wildlife interactions are increasing. At the 30th International Union of Game Biologists Congress held in Barcelona in early September 2011, in addition to two plenary presentations, 52 authors from 12 different countries and three continents presented 15 papers in the Interactions of Humans and Wildlife Session, three of which are included in this volume. To some extent, all the papers reflected the inherent difficulty in solving the complex problems caused either by rapidly increasing species that begin to inhabit urban and agricultural areas in numbers not seen previously (e.g. coyotes, Canis latrans, inhabiting big cities; wild boar, Sus scrofa, across western Europe; wood pigeons, Columba palumbus, in France), or species whose populations are threatened by human activities (e.g., Eurasian Lynx, Lynx lynx, in the Czech Republic). Some papers addressed the contentious issue of predator control (e.g., gamebirds in Great Britain), while others presented data regarding how human activities influenced animal behavior (e.g., pink footed geese, Anser brachyrhynchus; and red deer, Cervus elaphus, in Germany). The papers presented at the congress show how human activities affect the distributions and dynamics of wildlife populations and also change the behavior of some species. Wildlife causes social and economic conflicts by damaging agricultural and forest resources, bringing about traffic collisions, and creating problems for residents in urban areas; while many are increasingly distant from nature and may not accept the presence of wildlife others may actively encourage the presence of wild animals. The first paper in this volume, by Cahill et al. (2012), analyzes the management challenges of the increasing abundance of wild boar in the peri–urban area of Barcelona. This conflict has arisen in other large cities in Europe and elsewhere. The presence of the species causes problems for many residents, to such an extent that it is considered a pest in these areas. Wild boar habituation has not only been facilitated by population expansion, but also by the attitudes of some citizens who encourage their presence by direct feeding. This leads to wild boar behavior modification and also promotes an increase in the fertility rate of habituated females, which are significantly heavier than non–habituated females. Public attitudes regarding the species and harvesting methods (at present most specimens are removed by live capture and subsequently sacrificed) are highlighted as one of the key factors in the management of the conflict. The second paper provides an example of how the distribution of irrigated croplands influences wild boar roadkills in NW Spain (Colino–Rabanal et al., 2012). By modeling the spatial distribution of wild boar collisions with vehicles and using generalized additive models based on GIS, the authors show that the number of roadkills is higher in maize croplands than in forested areas. This factor is the main explanatory variable in the model. The paper provides an excellent example of how the synergies of diverse human elements in the landscape (maize croplands and roads in this case) affect the location and dimensions of these types of conflicts.

Carme Rosell, MINUARTIA, Dept. de Biologia Animal, Univ. de Barcelona. E–mail: crosell@minuartia.com Francesc Llimona, Estació Biològica can Balasc, Parc de Collserola. E–mail: fllimona@parccollserola.net ISSN: 1578–665X

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The third and final paper, by Belotti et al. (2012), addresses the effects of tourism on Eurasian lynx movements and prey usage at Šumava National Park in the Czech Republic. The monitoring of 5 GPS–collared lynxes and analyses of data regarding habitat features suggests that human disturbance (proximity of roads and tourist trails) can modify the presence of lynxes during the day close to the site where they have hidden a prey item, such as an ungulate, that can provide them with food for several days. In such cases, adequate management of tourism development must involve a commitment to species conservation. The analyses and understanding of all these phenomena and the design of successful wildlife management strategies and techniques used to mitigate the conflicts require a good knowledge base that considers information both about wildlife and human attitudes. The papers presented stress the importance of spatial analyses of the interactions and their relationship with landscape features and the location of human activities. Species distribution and abundance are related to important habitat variables such as provision of shelter, food, comfortable spaces, and an appropriate climate. Therefore, it is essential to analyze these data adequately to predict where conflicts are most likely to arise and to design successful mitigation strategies. The second key factor for adequate management of human–wildlife interactions is to monitor system change. An analysis of the variety of data on population dynamics, hunting, wildlife collisions, and wildlife presence in urban areas would provide a basis for adaptive management. In this respect, in the plenary session, Steve Redpath mentioned the importance of the wildlife biologist’s attitude when interpreting and drawing conclusions from recorded data and stressed the importance of conducting clear, relevant, and transparent science for participants involved in the management decision process, which often involves a high number of stakeholders. All of the papers addressing the problems associated with human wildlife interactions were characterized by a common theme. Regardless of the specific nature of the problem, the public was generally divided on how the problem should be addressed. A particularly sensitive theme was that of population control methods, especially when conflicts are located in peri–urban areas. Several presenters acknowledged that public participation was necessary if a solution was to be reached. Some suggested, as have other authors (Heydon et al., 2010), that a legislative framework may be needed to reconcile human and wildlife interests. However, each problem that was presented appeared to involve multiple stakeholders with different opinions. Solving these kinds of problems is not trivial. Social factors strongly influence perceptions of human–wildlife conflicts but the methods used to mitigate these conflicts often take into account technical aspects but not people’s attitudes. A new, more innovative and interdisciplinary approach to mitigation is needed to allow us 'to move from conflict towards coexistence' (Dickman, 2010). Other authors also mentioned the importance of planning interventions that optimize the participation of experts, policy makers, and affected communities and include the explicit, systematic, and participatory evaluation of the costs and benefits of alternative interventions (Treves et al., 2009). One technique that has been used to solve problems like these is termed Structured Decision Making (SDM). This technique was developed by the U.S. Geological Survey and the U.S. Fish and Wildlife Service. As described by Runge et al. (2009), the process is 'a formal application of common sense for situations too complex for the informal use of common sense', and provides a rational framework and techniques to aid in prescriptive decision making. Fundamentally, the process entails defining a problem, deciding upon the objectives, considering the alternative actions and the consequences for each, using the available science to develop a model (the plan), and then making the decision how to implement (Runge et al., 2009). Although complex, SDM uses a facilitator to guide stakeholders through the process to reach a mutually agreed–upon plan of action. It is clear that human–wildlife interactions are inherently complex because many stakeholders are usually involved. A rational approach that incorporates all interested parties would seem to be a productive way of solving these kinds of problems. References Cahill, S., Llimona, F., Cabañeros, L. & Calomardo, F., 2012. The increasing dilemma of wild boar (Sus scrofa) habituation to urban areas: traits from Collserola Park (Barcelona) and comparison with this problem in other cities. Animal Biodiversity and Conservation, 35.2: 221–233. Colino–Rabanal, V. J., Bosch, J., Muñoz, Mª J. & Peris, S. J., 2012. Influence of new irrigation croplands on wild boar (Sus scrofa) roadkills in NW Spain. Animal Biodiversity and Conservation, 35.2: 247–252. Belotti, E., Heurich, M., Kreisinger, J., Šustr, P. & Bufka, L., 2012. Prey usage by the Eurasian lynx (Lynx lynx): influence of human activity. Animal Biodiversity and Conservation, 35.2: 235–246. Dickman, J. A., 2010. Complexities of conflict: the importance of considering social factors for effectively resolving human–wildlife conflict. Animal Conservation, 13.5: 458–466. Heydon, M. J., 2010. Wildlife conflict resolution: a review of problems, solutions and regulation in England. Wildlife Research, 37.8: 731–748. Runge, M. C., Cochrane, J. F., Converse, S. J., Szymanski, J. A., Smith, D. R., Lyons, J. E., Eaton, M. J., Matz, A., Barrett, P., Nichols, J. D., Parkin, M.J., Motivans, K. & Brewer, D. C. 2009. Introduction to structured decision making, 5th edition. U. S. Fish and Wildlife Service, National Conservation Training Center, Shepherdstown, West Virginia, USA. Treves, A., Wallace, R. B. & White, S., 2009. Participatory planning of interventions to mitigate human–wildlife conflicts. Conservation Biology, 23.6: 1577–1587.

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Characteristics of wild boar (Sus scrofa) habituation to urban areas in the Collserola Natural Park (Barcelona) and comparison with other locations S. Cahill, F. Llimona, L. Cabañeros & F. Calomardo Cahill, S., Llimona, F., Cabañeros, L. & Calomardo, F., 2012. Characteristics of wild boar (Sus scrofa) habituation to urban areas in the Collserola Natural Park (Barcelona) and comparison with other locations. Animal Biodiversity and Conservation, 35.2: 221–233. Abstract Characteristics of wild boar (Sus scrofa) habituation to urban areas in the Collserola Natural Park (Barcelona) and comparison with other locations.— The parallel growth of urban areas and wild boar populations in recent years has increased the presence of this species around cities and in suburban areas, often leading to conflict with local people. In the Collserola Natural Park, situated within the metropolitan area of Barcelona, wild boar have become habituated to humans and urban settings because of direct feeding by local residents. Their attraction to these areas due to an abundance of anthropogenic food sources is especially strong during the warmer summer season when foraging conditions are poorer in their natural woodland habitat; the number of captures of habituated wild boar in peri–urban areas is significantly correlated with mean monthly temperatures. Habituated boar are primarily matriarchal groups, whereas adult and sub–adult (>1 year) males are significantly less represented than in non–habituated boars. In Collserola, habituated sub–adult and adult females are significantly heavier than their non–habituated counterparts and these weight differences increase with age; in the > 3 year–old age class they may be 35% heavier. Conflicts generated by the presence of wild boar in peri–urban areas are complex, and the responses by authorities are similarly diverse and often exacerbated by ambivalent public attitudes, both towards wild boar presence and applied mitigation measures. By 2010, at least 44 cities in 15 countries had reported problems of some kind relating to the presence of wild boar or feral pigs. Key words: Habituation, Human–wildlife conflict, Sus scrofa, Urbanisation, Wild boar. Resumen Características de la habituación de jabalíes (Sus scrofa) a las áreas urbanas en el Parque Natural de la Sierra de Collserola y comparación con otros lugares.— El crecimiento paralelo de las zonas urbanas y de las poblaciones de jabalíes durante los años recientes ha significado un aumento de la presencia de esta especie en las proximidades de las ciudades y de las áreas suburbanas donde a menudo representan una fuente de conflicto con las personas. En el Parque Natural de la Sierra de Collserola, situado en el área metropolitana de Barcelona, el jabalí se ha habituado a las personas y a las áreas urbanas como consecuencia de la alimentación directa por parte de vecinos. Su atracción a dichas áreas debido a una abundancia de alimento de origen antropogénico es especialmente fuerte durante los veranos cálidos cuando las condiciones tróficas son peores en su hábitat forestal natural; el número de capturas de jabalíes habituados en áreas periurbanas está significativamente correlacionado con las temperaturas medias mensuales. Los jabalíes habituados son principalmente grupos matriarcales, mientras que los machos adultos y subadultos (> 1 año) están significativamente menos representados, a diferencia de lo que se observa en los no habituados. En Collserola, las hembras adultas y subadultas habituadas pesan significativamente más que las hembras no habituadas y las diferencias de peso entre ellas incrementan a mayor edad; las > 3 años pueden pesar un 35% más. Los conflictos generados por la presencia de jabalíes en áreas periurbanas son complejos, y las respuestas por parte de las autoridades son también diversas y a menudo exacerbadas por unas actitudes ambivalentes por parte del público, tanto en lo que se refiere a la presencia del jabalí como a las medidas de mitigación aplicadas. Hasta el 2010, por lo menos 44 ciudades de 15 países habían registrado problemas de algún tipo relacionado con la presencia de jabalíes, o cerdos asilvestrados. ISSN: 1578–665X

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Palabras clave: Habituación, Conflicto, Sus scrofa, Urbanización, Jabalí. Received: 2 I 12; Conditional acceptance: 19 III 12; Final acceptance: 27 IV 12 S. Cahill, F. Llimona, L. Cabañeros & F. Calomardo, Consorci del Parc Natural de la Serra de Collserola, ctra. de l’Església 92, E–08017 Barcelona, España (Spain). Corresponding author: S. Cahill. E–mail: scahill@parccollserola.net

Animal Biodiversity and Conservation 35.2 (2012)

Introduction The rapid expansion of urban areas means that metropolitan landscapes are becoming increasingly significant from the perspective of wildlife ecology. Apart from the direct loss of habitat which occurs as a result of urban sprawl, there has been a dramatic increase in the contact zone between urban areas and wildlife habitat, often referred to as the wildland–urban interface (Radeloff et al., 2008). Such interfaces usually occupy more land than the built urban/suburban area itself, thus affecting wildlife habitat and human–wildlife interactions over an ever increasing surface area (Zhang et al., 2008). Although many species of wildlife are negatively affected by urbanisation processes (McDonald et al., 2008), others, particularly so–called generalist species, are capable of successfully exploiting habitat at the wildland–urban interface and even thrive in highly artificial urban greenspace and gardens. In recent times, racoons (Procyon lotor), fox squirrels (Sciurus niger), and coyotes (Canis latrans) have become established in metropolitan areas of Canada and the United States (Gehrt, 2007), while red fox (Vulpes vulpes) frequently occupy urban areas (Harris & Smith, 1987; König, 2008). Likewise, wild boar (Sus scrofa) populations are currently expanding worldwide both in distributional range and in numbers in many countries where they are present, either as a native or non–native species. Wild boar population dynamics are often dictated by the abundance or scarcity of pulsed resources, such as mast seeding by oak (Quercus spp.) and beech (Fagus spp.) trees (Bieber & Ruf, 2005). Nevertheless, this species has benefitted from changes in traditional agro–silvo–pastoral landscapes which have removed the limitations that hitherto existed for population growth. Among other factors, the loss of predators, the intensification of agricultural practices, supplementary feeding, the deliberate release by hunters, and even global warming (Geisser & Reyer, 2005) all have contributed to increased abundance of wild boar. Previously confined mainly to rural, forested, mountainous, and similar natural areas with low human presence, in recent years wild boar have become increasingly habituated to urban areas (Kotulski & König, 2008). Among wildlife, habituation is defined as the loss of fear response to the presence of humans after repeated, non–consequential encounters (Herrero et al., 2005; McNay, 2002; Wieczorek–Hudenko & Decker, 2008). However, conflicts arise when the species’ presence overlaps both in time and space with human activity and their activities become an annoyance to residents (Loker & Decker 1998). Such conflicts are varied, ranging from general nuisance to more serious issues such as disease transmission, increased risk of traffic accidents, or even attacks on humans (Hubbard & Nielsen, 2009; Storm et al., 2007; Timm et al., 2004; White & Gehrt, 2009). Wild boar have become very abundant in the province of Barcelona (NE Spain) over the last two decades, with an eight–fold increase in annual hunting bag returns between the 1986 and 2005 seasons (fig. 1). This increase has coincided with a rapid rise in the human population in the province, which has increased

223

from just over 4.5 million to 5.5 million, and an intense period of sprawling urbanisation (fig. 1). However, hunting has been forbidden in the area covered by the Barcelona city municipality for more than 20 years, and groups of wild boar in the adjacent Collserola Natural Park (fig. 2) have become increasingly habituated to human presence over the last 10 years, especially in peri–urban areas located beside this and other nearby cities (fig. 3). Habituation of wild boar has largely been motivated by direct feeding by people and is also facilitated by the proximity of densely vegetated areas close to the city limits (Llimona et al., 2007). This habituation leads to frequent conflicts for neighbours, park managers, and city authorities due to damage caused to gardens and landscaped areas, as well as fear of attacks and collisions with vehicles. In this study we aimed to identify the main environmental factors associated with habituation incidence in the Collserola Natural Park (Barcelona, NE Spain), and to investigate potential differences in relation to sex, age, and weight between habituated and non–habituated wild boar. We describe the characteristics, consequences, and management implications of wild boar habituation to urban areas near the Park, and contextualise the increasing presence of wild boar and feral swine in peri–urban situations in other countries. Study area The Collserola Natural Park (41° 25' 52'' N, 2° 4' 45'' E) is a Natura 2000 site situated in the middle of the Barcelona Metropolitan Area (fig. 2). It comprises 36 municipalities with a population of 3.2 million people and a population density of over 5,000 inhabitants per km2 (INE 2011). Collserola occupies ~9,000 ha of mountainous (60–512 m a.s.l.) Mediterranean scrub and woodland (60%), with declining agriculture and intense urban and infrastructure pressure on its periphery. Wild boar numbers have increased in Collserola since the early 1980s, and the population has been estimated to be about 800 individuals (density ~11 boar/km2) based on hunting returns in recent years (Cahill et al., 2012). The hunting of wild boar is permitted in ~50% of available habitat, and on average about 100 wild boar are killed there each year by hunters (annual range of 61–192 captures for 2004–2011). Hunting occurs either via large battues (drive hunts with hunters placed at fixed positions and teams of dogs used to flush the boars) carried out between October and February, or with special permits authorised for damages throughout the year (fig. 4; Cahill et al., 2012). Further details on the study area and on the demography and biology of wild boar in Collserola may be found in Cahill et al. (2003) and Cahill & Llimona (2004), and on the specific peri–urban metropolitan context of Collserola’s wildlife in Llimona et al. (2005, 2007). Material and methods We define habituated wild boar in this study as individuals that are clearly accustomed to or indifferent to

5,600,000

Wild boar hunted in Barcelona province

5,400,000

10,000

New residential units built

10–11

08–09

Year/hunting season

06–07

86–87

4,400,000

04–05

2,000

02–03

4,600,000

00–01

4,000

98–99

4,800,000

96–97

6,000

94–95

5,000,000

92–93

8,000

90–91

5,200,000

88–89

Human population of Barcelona province

12,000

Human population

Wild boar hunted per season/ residential construction in Catalonia (housing/apartment units)

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Fig. 1. Trends in human population growth (left ordinate) and the number of wild boar hunted per season in the province of Barcelona from 1986 to 2010, and residential construction in Catalonia since 1990 (right ordinate). Data from INE (2011), the Territorial Service of the Government of Catalonia, and http:// www20.gencat.cat/portal/site/ptop. The vertical arrow marks the year in which the systematic live capture of habituated wild boar began in Collserola. Fig. 1. Tendencias de crecimiento de la población humana (eje de ordenadas izquierdo) y el número de jabalíes cazados por temporada en la provincia de Barcelona entre 1986 y 2010, y de construcción de viviendas en Cataluña desde 1990 (eje de coordenadas derecho). Datos del INE (2011), los Servicios Territoriales de la Generalidad de Cataluña y http://www20.gencat.cat/portal/site/ptop. La flecha vertical indica el año en que se inició la captura sistemática en vivo de jabalíes habituados en Collserola.

human presence and usually do not flee from people. Live capture of habituated wild boar in urban areas of Collserola has been carried out on a regular basis by park wardens using tranquiliser darts since 2004 to address complaints either from residents, local police, or other authorities. Wardens dart habituated wild boar during daylight at close range (~5–15 m) with Zoletil® 100 (VIRBAC S.A., Spain), an anaesthetic mixture of tiletamine and zolazepam (see Fournier et al. (1995) for its use on wild boar), applied using a JM Special rifle (DAN–INJECT ApS, Denmark). It is usually possible to capture most or all of the members of sounders located in urban areas on a first attempt, and if not, they are captured on subsequent visits given that groups quickly return to the same site in the urban area. At present, most wild boar captured are subsequently euthanized. We gathered biometric data from captured boar and estimated age class based on tooth eruption sequences (Monaco et al., 2003). We also collected comparable data from wild boar mortality from different causes (e.g., hunting and road kills). We correlated the number of wild boar captured in urban areas with mean monthly maximum temperatures and means of total monthly rainfall using values from local meteorological data available online (http://www.

fabra.cat/meteo/dades/dades.html). In order to obtain information regarding foraging conditions in natural areas of Collserola, we gathered relevant data during seasonal plot surveys carried out between September 1998 and December 2004. We assessed acorn availability, soil rooting conditions, and wild boar rooting activity at 27 fixed 3 m x 3 m plots located in different woodland areas within Collserola Natural Park. We evaluated soil rooting conditions through calculation of a simple index based on soil humidity and compactedness both at the surface layer and at 10 cm depth, and this index takes values between 0 (poorest conditions) and 1 (optimal conditions). The specific methodology applied in surveying these seasonal plots is detailed further in Cahill et al. (2003) and Cahill & Llimona (2004). We used official data on motorway traffic through Collserola as an indicator of monthly human presence for 1997–2007 from http://www20. gencat.cat/portal/site/ptop. For statistical analysis, we used Chi–squared, t–tests, and one–way ANOVA to evaluate comparisons of proportions and means and Spearman’s Rank correlation coefficient for the analysis of linear correlations between variables (Zar, 1984). Where mean values are quoted these are provided with their corresponding standard errors.

Animal Biodiversity and Conservation 35.2 (2012)

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Spain

Parc de Collserola Barcelona

Fig. 2. Location of the Collserola Natural Park beside the city of Barcelona. Fig. 2. Situación del Parque Natural de la Sierra de Collserola al lado de la ciudad de Barcelona.

Barcelona

Fig. 3. Locations of incidents concerning the presence of wild boar in peri–urban areas of the Collserola Natural Park and its surroundings. Fig. 3. Localización de los incidentes relacionados con la presencia de jabalíes en áreas periurbanas del Parque Natural de la Sierra de Collserola y de sus alrededores.

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Number of individuals

150 120

Live capture using anaesthetic darts Killed in battues/nocturnal waits

90 60 30 0 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010

Fig. 4. Number of non–habituated wild boar killed by hunters (n) and habituated wild boar (¡) captured by park wardens using tranquilizer darts in the Collserola Natural Park each year since 2000. (Data are calendar year totals.) Fig. 4. Número de jabalíes no habituados capturados por cazadores (n) y de jabalíes habituados (¡) capturados por guardas del parque mediante dardos anestésicos en el Parque Natural de la Sierra de Collserola cada año desde el 2000. (Los datos son totales para cada año.)

Table 1. Seasonal mean values of the rooting suitability index, the percentage of plot rooted, and acorn density at fixed plots monitored in Collserola between September 1998 and December 2004, with corresponding p–values of Tukey HSD post–hoc multiple comparisons. Tabla 1. Valores medios mensuales del índice de condiciones para hozar, del porcentaje de parcela hozada y de la densidad de bellotas en parcelas fijas muestreadas en Collserola entre septiembre de 1998 y diciembre de 2004, y los valores correspondientes de p de comparaciones múltiples post–hoc realizadas mediante la prueba HSD de Tukey. Season

Mean

S.E.

N plots

Summer

Autumn

Winter

Rooting suitability index Spring

0.389

0.015

135

p = 0.000

Summer

0.249

0.017

194

–

p = 0.000

Autumn

0.466

0.011

216

–

p = 0.840

Winter

0.481

0.009

160

Spring

1.259

0.568

135

p = 0.721

p = 0.917

p = 0.257

Summer

0.699

0.263

194

–

p = 0.226

p = 0.010

Autumn

1.597

0.290

216

–

p = 0.514

Winter

2.280

0.373

160

Spring

2.655

1.028

135

p = 0.735

p = 0.939

p = 0.098

Summer

0.669

0.300

194

–

p = 0.274

p = 0.002

–

p = 0.199

–

Plot rooted (%)

–

Acorn density (per m2)

Autumn

3.747

1.223

216

Winter

7.304

2.153

160

–

Animal Biodiversity and Conservation 35.2 (2012)

100.0

Wild boar captured Mean precipitation Mean temperature

50.0

90.0 80.0 70.0

40.0

60.0 30.0

50.0 40.0

20.0

30.0 20.0

10.0

10.0 0.0

J J Months

Mean montly precipitation

Wild boar live–captured/ mean monthly temperature

60.0

227

0.0

Fig. 5. Total number of wild boar live–captured per month in peri–urban areas of Collserola in relation to mean monthly temperature (in ºC, solid line and squares) and mean monthly precipitation (in l/m2, dotted line and circles) between 2004 and 2010. Fig. 5. Número total de jabalíes capturados cada mes en las áreas periurbanas de Collserola en relación con las temperaturas medias mensuales (en ºC, línea continua con cuadrados) y precipitación media mensual (en l/m2, línea discontinua con círculos) entre 2004 y 2010.

Results The mean index of soil suitability for rooting activity was significantly different between seasons (F3,696 = 4.28, p < 0.0001), with poorest conditions in summer and best conditions in winter (table 1). The mean percentage surface area of plots that was rooted by wild boar also varied between seasons (F3,696 = 3.23, p < 0.05), again being lowest in summer and highest in winter (table 1). Likewise, mean acorn density varied between seasons (F3,696 = 4.33, p < 0.005), being lowest in summer and highest in winter (table 1). A total of 293 habituated wild boar were captured live and removed from urban areas of Collserola between 2004 and 2010 (fig. 4), representing a mean of 42 ± 5,9 animals captured per year (range: 19–66). Live captures of habituated wild boar tended to be concentrated during the warmer months of the year, especially between May and October (fig. 5), and the monthly total of such captures was significantly correlated with mean monthly temperatures (rs = 0.771, p < 0.005, n = 12, data from 2004–2010 inclusive). In contrast, there was no correlation between the number of incidents and mean monthly precipitation during this same period (fig. 5, rs = 0.133, p = 0.68, n = 12). Fewer captures were recorded in August, despite the high mean temperature for this month (fig. 5). Lower human presence in the Collserola area during August is indicated by data from the E–9 toll motorway which runs through the park, with 47% less traffic than in the

other months of the year (F1,11 = 6.371, p < 0.0001: ~16,000 vehicles/day in August and minimum–maximum range for the remaining months between ~28,000 in January and ~33,000 vehicles/day in November). In contrast, there was a sharp increase in wild boar captures during the month of September, which may reflect residents’ return from holidays. When total captures for the months of August and September (n = 74) are averaged, thus adjudicating 37 captures to each month, the overall relation between monthly temperature and captures of habituated wild boar is strengthened (rs = 0.912, p < 0.001, n = 12). Groups of habituated wild boar were comprised almost entirely of adult females accompanied by sub– adults and piglets, whereas adult male boars were less frequently encountered in urban areas (fig. 6). Among habituated males, adults and sub–adults (over 1 year old) comprised a significantly lower proportion (31.2%) of individuals than those in an accumulated sample of non–habituated males in which 64.3% were more than 1 year old (c2 = 15.86, p < 0.001, n = 64 habituated and 84 non–habituated individuals). In the case of females, the proportions of individuals over 1 year old among habituated (48.9%) and non–habituated (62.4%) wild boar were not significantly different (c2 = 3.21, p = 0.07, n = 90 habituated and 85 non–habituated). Habituated wild boar females were significantly heavier than non–habituated females in all age classes > 1 year old, and weight differences between them increased in older age classes (table 2). On

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Table 2. Mean weights (kg) of different age classes of habituated and non–habituated male and female wild boar in Collserola. T–test values and corresponding significance levels (p value) are indicated for appropriate t–test comparisons of means between habituated and non–habituated animals. Sample sizes are indicated in brackets following means and standard errors. Tabla 2. Pesos (kg) medios de las distintas clases de edad de jabalíes machos y hembras habituados y no habituados en Collserola. Se indican los valores de test t y sus niveles correspondientes de significado (valores de p) para las comparaciones de promedias entre animales habituados y no habituados. El tamaño de las muestras se indica entre paréntesis después de los valores promedios y sus errores estándar. Males

Habituated

Females Non–habituated

Habituated

Non–habituated

0–6 months 15.66

2.04 (24) 13.44

t = –0.836

1.41 (18)

14.15

1.93 (20) 15.16

p = 0.408 t = 0.485

1.06 (26)

p = 0.630

6–12 months 37.19

2.85 (9) 30.98

t = –2.014

1.54 (11)

33.44

2.35 (20) 24.58

p = 0.059 t = –1.979

2.04 (6)

p = 0.059

1–2 years 53.63

3.27 (16) 48.02

t = –1.396

2.18 (40)

53.99

2.60 (27) 42.77

p = 0.168 t = –3.428

1.96 (26)

p = 0.001

2–3 years 84.03

4.92 (3) 71.71

t = –1.02

7.42 (7)

65.47

3.62 (11) 51.75

p = 0.338 t = – 3.657

1.60 (13)

p = 0.001

> 3 years 85.30

7.10 (2) 80.25

t = –0.375

7.14 (6)

83.34

4.75 (7) 61.78

p = 0.720 t = –3.507

average, habituated females were 26.2%, 26.5% and 34.9% heavier than non–habituated individuals in the 1–2 year, 2–3 year and 3 year+ age classes, respectively. Habituated female wild boar that were 6–12 months old were also heavier than non–habituated females, although this weight difference was marginally non–significant, while there was no significant difference between them in the 0–6 month old age class (table 2). Habituated males also weighed more than non–habituated males in all age classes. However, unlike females, these differences were not statistically significant, being only marginally non–significant in the 6–12 month age class. Discussion Several factors have coincided in facilitating the initial habituation of wild boar to urban areas and human presence in Collserola. In Mediterranean areas, and specifically as is shown in this study, food is scarcer in natural woodland habitat during the months of summer drought, mainly because foraging conditions are poorer

3.71 (13)

p = 0.003

as a result of the hardening of the soil. Also, during this period there is a scarcity of forest food sources, such as oak mast, which play an important role in wild boar demographics (Massei et al., 1996, 1997; Cahill & Llimona, 2004). Indeed, during the summer nocturnal feeding activity of wild boar in the Collserola mountains is concentrated on more humid, generally northern facing slopes, and in valley ravines and by streams, whereas feeding activity is almost non–existent during this season on the drier forest slopes that cover most of the surface area of the park (Cahill et al., 2003). Thus, wild boar might be expected to venture into peri–urban areas in search of anthropogenic food sources during summer if food is scarce in adjacent woodland areas. In this regard, mean monthly temperature appears to be a good correlate of habituation incidence in Collserola, unlike precipitation which is much more variable and irregular. Anthropogenic food sources in peri–urban areas are varied and can be abundant. In addition to direct feeding, unintentional (indirect) feeding is also important, either through food left out for domestic pets or discarded rubbish, and also via irrigated lawns, gardens, and other landscaped areas such as golf courses or ceme-

Animal Biodiversity and Conservation 35.2 (2012)

100 Percentage of cases (%)

Males

80 70 60 50

8 48

13 56

40 14

30 20

10 Yes (n = 64)

No (n = 84)

100 Percentage of cases (%)

Females

229

60 50 40 30 20

Age class > 3 years

2–3 years

1–2 years 6–12 months

10 0

Yes (n = 90) No (n = 85) Habituated wild boar

0–6 months

Fig. 6. Proportions of age classes represented among habituated and non–habituated male and female wild boars in the Collserola Natural Park. Data from 2004 to 2008. Fig. 6. Proporción de clases de edad representadas entre jabalíes habituados y no habituados de ambos sexos en el Parque Natural de la Sierra de Collserola. Datos del período 2004 a 2008.

teries. Such feeding opportunities encourage daytime activity and subsequently a loss of fear of people. As a consequence, wild boar behaviour surpasses mere habituation, which is defined more by indifference (Wieczorek–Hudenko & Decker, 2008). Indeed, boars are positively attracted to peri–urban areas, and in Collserola for example they frequently root up areas of grass in city parks and turn over large domestic rubbish bins to obtain food. The greater mean weight of habituated female wild boar compared to non–habituated individuals here evidences the importance of such resources for this species. Lactating sows have high energy requirements, and protein can also be a limiting nutrient for milk production (Barrett, 1978). In his study on the feral hog at the Dye Creek Ranch in California, Barrett (1978) found that most lactating feral sows showed poor body condition

during summer, and that piglet mortality was greatest during this period. Also, starvation was aggravated during poor mast periods and it increased the vulnerability of mortality due to causes such as accidents or predation (op. cit.). As such, it is important to emphasize that Mediterranean–type drought conditions can limit the reproductive performance of sows (Fernández–Llario & Carranza, 2000), and consequently the attraction towards readily accessible anthropogenic food sources is strong. For example, in the Mediterranean climate of California, Barrett (1982) found that during the critical hot dry summers, feral hogs that foraged in irrigated pastures of the Sacramento Valley showed greater reproductive success, higher growth rates, and lower juvenile mortality than hogs which remained in the nearby mountain foothills characterised by oak–dominated (Quercus spp.) habitats. Similarly, any alleviation of

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the drought–induced depression of reproductive performance through artificial feeding will also inevitably result in greater densities of wild boar in peri–urban Mediterranean areas. To date, few studies have investigated the presence of wild boar in or close to urban areas, and most of these have dealt with aspects relating to the incidence of zoonoses (Jansen et al., 2007; Nidaira et al., 2007; Schielke et al., 2009). Wild boar in urban areas are more than just a nuisance and may pose risks of disease transmission to humans (Meng et al., 2009). For example, some urban areas now known to have problems with wild boar are also located adjacent to large standing water bodies, as is the case in Berlin, Singapore, or Florida, where contamination of water may increase the risk of infection by Escherichia coli (Jay et al., 2007) or Leptospira sp. Stauffer (2010) reported that 22.9% of 192 people who had participated in an adventure race in Florida that involved swimming in open waters presented symptoms of Leptospirosis, with three of them requiring hospitalisation. Florida has a very large feral boar population, estimated at > 500,000 animals (Giuliano, 2011). Their distribution overlaps vast extensive residential urban developments in the midst of wetlands. Proximity of wild boar to domestic pets could also increase the risk of Toxoplasmosis gondii in relation to cats (Richomme et al., 2010), or Echinococcus granulosus/multilocularis (Boucher et al., 2005; Martín–Hernando et al., 2008) in relation to dogs. It is therefore important that the authorities are aware of such potential risks, though not necessarily alarmed, given that some of these risks already exist in peri–urban areas due to other wildlife such as rodents. Although few data are available in the scientific literature on the extent of wild boar presence in urban areas, there are frequent reports in media sources regarding this problem in various cities. We undertook a non– exhaustive search using the following keywords with Google: 'wild boar', 'wild/feral pig', 'Sus scrofa', 'jabalí', 'sanglier', 'wildschwein', 'cinghiale', in combination with 'city', 'town', 'street', 'urban', 'neighbourhood', 'residential', 'ciudad', 'ville', 'stadt' and 'città'. We found that, up until 2010, at least 44 cities or towns in 15 countries (Belgium, China, France, Germany, India, Israel, Italy, Japan, Korea, Poland, Singapore, Spain, United Kingdom, United States and Romania) had reports of incidents concerning wild boar/pigs. These figures are probably conservative given the limitations of this search (e.g., no searches were done in languages such as Russian, Chinese, or Japanese). Of these cities, 36.4% had experienced just one or two incidents, while the majority (63.6%) had already reported several or many cases. Many cities reporting problems concerning habituated wild boar, such as Genoa (Italy), Haifa (Israel), or San José, California (U.S.), are also located in regions with a Mediterranean or a subtropical climate, and appear to be characterised by similar patterns of dry summers combined with anthropogenic food–rich peri–urban landscapes. Nevertheless, cases of wild boar habituation are also found in cooler, temperate regions (e.g., Germany, Poland, UK, Japan) where snow cover or frozen soil can impede rooting activity in winter in a similar way to drought conditions in summer in Mediterranean areas.

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Food scarcity could indeed be a powerful driving force behind wild boar habituation, but other factors may be important. Despite important biogeographic differences between cities that have problems with habituation, most are commonly characterised by narrow gradients at the urban–rural interface, often typified by direct contact between urban and forested or densely vegetated areas, as for example in Kobe (Japan) or Berlin (Germany), where frequent problems occur with wild boar (Kobe City, 2012; Kotulski & König, 2008). Habituation in Collserola has coincided with a period of rapid growth in the wild boar population, which possibly saturated available natural habitat and thereby led to immigration into surrounding peri–urban areas. However, this same period was also characterised by important urban expansion in the area (fig. 1). Under this situation, wild boar populations and sprawling urbanisation were on a collision course which materialised as initial cases of habituation during the early 2000s, and later intensified as chronic conflicts requiring specific management strategies (Cahill et al., 2012). Landscape characteristics were shown to have a significant influence on the location of incidents relating to habituated wild boar in Collserola (Llimona et al., 2007). Contrast–weighted edge density values for sharp contrast gradients between urban and densely vegetated habitat, such as woodland, were significantly higher at known incident locations than in random locations. However, no differences were found in the case of soft gradient edge values corresponding to urban areas in contact with more openly vegetated habitat (op. cit.). Similarly, other parts of Spain that have experienced rapid urban expansion, such as Las Rozas (Madrid), have also recorded cases of habituation by groups of wild boar, especially during or after drought periods (López et al., 2010). Nevertheless, it is clearly the availability of anthropogenic food sources that attracts wild boar to peri–urban areas. Benign attitudes of urban residents towards wild boar have facilitated their habituation, either directly encouraging their presence by intentional feeding, or simply through indifference (habituation of people to boars). However, for some people the presence of wild boar is a nuisance because of the damage they ultimately cause to gardens, golf courses and , parks, as well as their rooting in rubbish bins and the spreading of garbage. Wild boar can be a danger to traffic on city streets, and they occasionally cause significant disruption on major roads around Collserola. Some residents are fearful of possible attacks on children or pets, and although few attacks are reported in relation to habituated wild boar in urban areas, non–habituated wild boar can be dangerous to people and are known to cause injuries and even fatalities in rural areas of northern and central India (Chauhan et al., 2009). As such, not all residents look favourably on the presence of wild boar in urban areas. In Berlin, wild boar have been frequenting peri–urban areas for the past two decades (Georgii et al., 1991) and as their numbers have increased sharply they are now considered as a serious nuisance. Kotulski & König (2008) carried out a survey on attitudes of Berliners towards wild boar

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and found that 23% of residents objected to their presence, 37% were in favour and 36% had ambivalent opinions. Despite the fact that 44% of people believed the numbers of wild boar should be reduced, 67% of these were against lethal removal methods. Massei et al. (2011) gave an extensive review of the diverse options currently available to control populations of wild boar/pigs. However, the ambivalent opinions of residents complicate the implementation of efficient management strategies for dealing with this species in urban areas. In cities where problems have arisen with habituated wild boar, implemented management options range from hunting with firearms to removal using live capture (darting and cage–traps), to public awareness campaigns, to doing nothing. In Collserola for example, live capture using tranquiliser darts is currently the usual method of removing habituated wild boar from urban areas. Recently, however, trial attempts were made by the authorities responsible for hunting to kill habituated wild boar in peri–urban areas using expert bow hunters equipped with modern bows and arrows. However, within less than a week of announcing this control method, it had to be withdrawn due to opposition pressure from the public, highlighting the difficulties in implementing apparently technically valid control options. Although archery is considered the preferred hunting method for controlling deer in urban areas of some U.S. states, such as Connecticut (Kilpatrick et al., 2002; Kilpatrick & LaBonte, 2007), its use in hunting in Europe is more restricted, and perhaps less culturally accepted by urban dwellers. In a study of residents’ acceptance of solutions to wildlife conflicts, invasive and lethal solutions were in fact more highly accepted than had been expected by wildlife managers (Loker et al., 1999). Nevertheless, suburban residents were also more likely to accept non–lethal management actions (op. cit.). In early cases in Collserola, attempts were made to relocate habituated wild boar several kilometres away from the urban areas where they had been causing problems. However, radiotracking of these individuals showed that they quickly returned to the exact same places where they had been captured. Indeed, previous research on translocating feral hogs in California has shown that groups need to be moved to suitable habitat at least 15–20 km away in order to avoid their returning (Barrett, 1978; Lewis, 1966). Subsequently, some problem animals from Collserola were sent to enclosed private estates, but now most habituated wild boar captured using tranquilizer darts are subsequently euthanized, although again, public opinion is often against this. Drive hunting (with dogs) of wild boar is carried out in forested areas of Collserola, and although it may be tempting to conclude that increased hunting reduces habituation incidence (fig. 4), in peri–urban areas such control methods may be impracticable, or they may meet with opposition from the public and land–owners (Storm et al., 2007). Given this situation, live–capture methods, either using tranquiliser darts or trap–cages, will continue to play an important role in removing specific problem individuals or sounders. It is important to stress in public relations that these methods are quick and humane in removing animals.

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Given the limitations of efforts to reduce human–wild boar conflicts solely through lethal methods aimed at population suppression (Honda & Kawauchi, 2011), specific alternative measures are also required to prevent and mitigate habituation in order to alleviate the growing incidence of conflicts in urban areas. In Collserola, emphasis is placed on prevention through public awareness campaigns. Specifically, public education is required on the eco–ethological consequences of feeding wild boar, as well as on the ultimate destiny of most habituated individuals (capture and euthanasia). Nevertheless, other preventive measures are also required: the possibility of fines has been introduced by city councils to dissuade direct feeding, and the landscaping of peri–urban green space needs to be adapted to make it less attractive to wild boar, for example by reducing the irrigation of grassy areas. Domestic rubbish bins should be designed so that they are inaccessible to wild boar, as should feeding points destined for domestic pets. Improved fencing is also necessary in many private gardens as well as suburban parks to impede wild boar access. Also, the maintenance of more open vegetation (e.g., native dry grasslands) adjacent to peri–urban areas may also help to reduce wild boar presence. Once again, public education is also required in order to gain maximum acceptance of such mitigation measures. Acknowledgements We thank all our colleagues, present and past, at Can Balasc, the Collserola Natural Park Biological Station, for their assistance in many aspects of our work on wild boar, and in particular to Susanna Carles and Álvaro Picañol. We especially thank the park wardens, Francisco Javier García, Àngel Mateo, Jordi Piera and Carles Sobrino for their intense work at the front line with wild boar in Collserola and all that this involves. Our thanks to the Cos d’Agents Rurals de la Generalitat de Catalunya for their assistance on occasions with the capture of habituated wild boar, and thanks also to Jordi Gràcia for his assistance with field work during earlier studies on wild boar ecology in Collserola. Thanks to Joan Roldan of Forestal Catalana, S. A. for providing data on wild boar hunting returns for the province of Barcelona. Finally, our thanks to the Associate Editor, to Dr. Dale Rollins of Texas Agrilife Research, and to another anonymous referee for their comments and improvements on an earlier draft of this manuscript. References Barrett, R. H., 1978. The feral hog on the Dye Creek Ranch, California. Hilgardia, 46: 283–355. – 1982. Habitat preferences of feral hogs, deer, and cattle on a sierra foothill range. Journal of Range Management, 35(3): 342–346. Bieber, C. & Ruf, T., 2005. Population dynamics in wild boar Sus scrofa: ecology, elasticity of growth rate and implications for the management of pulsed

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Influence of tourism and traffic on the Eurasian lynx hunting activity and daily movements E. Belotti, M. Heurich, J. Kreisinger, P. Šustr & L. Bufka

Belotti, E., Heurich, M., Kreisinger, J., Šustr, P. & Bufka, L., 2012. Influence of tourism and traffic on the Eurasian lynx hunting activity and daily movements. Animal Biodiversiy and Conservation, 35.2: 235–246. Abstract Influence of tourism and traffic on the Eurasian lynx hunting activity and daily movements.— Human presence influences survival of large carnivores in several ways and even outdoor activities can be a source of disturbance. As ungulate prey provide the Eurasian lynx (Lynx lynx) with food for several nights and the pattern of lynx activity is mainly shaped by searching for and consuming large prey, the need to move decreases strongly while the prey is eaten. However, during the day, human activity may drive lynx to move to safe shelters and habitat features such as dense vegetation may increase tolerance. In the Bohemian Forest (Czech Republic), we found 116 prey killed by five GPS–collared lynxes. We tested whether the kill sites were located farther from roads or tourist trails than a set of randomly generated locations and whether presence of roads or tourist trails and habitat structure influenced the distance 'kill site to daytime resting sites'. At night, with low human activity, lynxes did not avoid roads and even selected the surroundings of tourist trails. The distance 'kill site to daytime resting sites' correlated negatively with presence of habitat concealment and distance to tourist trails, suggesting that outdoor activities may have to be considered in lynx management plans. Key words: Lynx lynx, Kill site, Resting site, Tourist trail, Paved road, Habitat structure. Resumen Influencia del turismo y del tráfico sobre la caza del lince boreal y sus desplazamientos diarios.— La presencia humana influye de diversas formas sobre la supervivencia de los grandes carnívoros, e incluso las actividades al aire libre pueden ser una fuente de perturbaciones. Dado que los ungulados son la presa que proporciona al lince boreal (Lynx lynx) alimento para varias noches, y que el patrón de la actividad del lince está diseñado principalmente para buscar y consumir presas de gran tamaño, la necesidad de desplazarse disminuye mucho mientras está devorando la presa. No obstante, durante el día, la actividad humana puede obligar al lince a desplazarse a refugios seguros, y las características del hábitat tales como una vegetación densa pueden aumentar su tolerancia. En los bosques de Bohemia (República Checa), hallamos 116 presas cazadas por cinco linces provistos de collares GPS. Estudiamos si los lugares de la matanza estaban situados más lejos de las carreteras o de los senderos turísticos que si los lugares hubieran sido elegidos al azar, y si la presencia de carreteras o senderos turísticos y la estructura del hábitat influían en la distancia “lugar de caza a zonas de descanso diurnas”. Por la noche, con una actividad humana baja, los linces no evitaban las carreteras e incluso elegían los alrededores de los senderos turísticos. La distancia “lugar de caza a zonas de descanso diurnas” estaba correlacionada negativamente con la presencia de hábitat críptico y con la distancia a los senderos turísticos, lo que sugiere que las actividades al aire libre podrían tomarse en consideración en los planes de gestión de los linces. Palabras clave: Lynx lynx, Lugar de caza, Lugar de reposo, Sendero turístico, Carretera pavimentada, Estructura del hábitat. Received: 11 I 12; Condtional acceptance: 19 III 12; Final acceptance: 9 V 12 E. Belotti & L. Bufka, Fac. of Forestry and Wood Sciences, Czech Univ. of Life Sciences Prague, Kamýcká 1176, CZ–16521 Prague 6, Czech Republic.– E. Belotti, L. Bufka & P. Šustr, Dept. of Research and Nature Protection, Šumava NP and PLA Administration, Sušická 399, CZ–34192 Kašperské Hory, Czech Republic.– M. Heurich, Bavarian Forest National Park, Freyunger Str. 2, D–94481 Grafenau, Germany.– J. Kreisinger, Dept. of Zoology, Fac. of Science, Charles Univ. in Prague, Viničná 7, CZ–12844 Prague 2, Czech Republic. Corresponding author: E. Belotti, E–mail: belotti@fld.czu.cz

ISSN: 1578–665X

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Introduction Since the 1950s, after almost two centuries of dramatic declines and extinctions, the populations of large carnivores are slowly recovering in several European countries (Linnell et al., 2001, 2005). This has been achieved by a change in people´s attitude towards these species and by the consequent adoption of favorable legislation, as a result of both spontaneous expansion and reintroduction programs (Boitani, 2000; Breitenmoser et al., 2000; Swenson et al., 2000). At least in Central Europe, nonetheless, only few areas can be still considered 'natural' or 'semi–natural' and most of the areas to which bear (Ursus arctos), Eurasian lynx (Lynx lynx) and wolf (Canis lupus) are returning include human–modified habitats, which can be quite different from their original habitats (Basille et al., 2009). Therefore, their ability to coexist with man is probably one of the most important factors influencing their long–term survival (Breitenmoser et al., 2000; Boitani, 2000; Basille et al., 2009). There are several human activities that can negatively affect the survival of large carnivores. For example, hunting, poaching (Jędrzejewska et al., 1996; Andrén et al., 2006) and road intensive usage causing vehicle collisions (Kaczensky et al., 2003; Andrén et al., 2006) have a direct effect, while forestry activity and road network development leading to habitat fragmentation (Theuerkauf et al., 2001; Huck et al., 2010) and game management influencing prey availability and distribution (e.g. Putman & Staines, 2004; Milner et al., 2007; Hothorn & Müller, 2010) can have an indirect influence. In spite of this, large carnivores in Europe still show a certain degree of tolerance to humans (Linnell et al., 2000; Theuerkauf et al., 2003; Bunnefeld et al., 2006). Studies have shown that they are able to permanently occupy areas with a low degree of urbanization (Basille et al., 2008, 2009) and survive, under certain conditions, even in areas with high human density (Linnell et al., 2001). On the other hand, in natural and semi–natural areas, generally associated with a low human density, there has been a noticeable increase in outdoor activities, especially in the last decades (Thiel et al., 2008; Balmford et al., 2009). Concerning the influence of such activities, an increasing number of studies (Burger & Gochfeld, 1998; Duchesne et al., 2000; Taylor & Knight, 2003; Dyck & Baydack, 2004; Thiel et al., 2007) seems to prove that nonlethal disturbance stimuli can produce the same effect as predation risk on the species fitness: they might induce an 'antipredator response' that has a cost to other activities (Frid & Dill, 2002). In addition, human hunters have represented a real threat for large carnivores over evolutionary time (Frid & Dill, 2002). Thus, in some cases, such as when people approach on foot, disturbance stimuli and true predatory stimuli may be indistinguishable from these animals’ perspective (Frid & Dill, 2002). Although literature about the effects of human disturbance on animal behavior is quite rich and several authors have studied the effects of human activity in general on large carnivores (Amstrup et al., 1993; Thurber et al., 1994; Kerley et al., 2002; Theuerkauf

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et al., 2003; Bunnefeld et al., 2006; Kolowski & Holekamp, 2009), to date only a few studies (e.g. Goodrich & Berger, 1994, reviewed in Linnell et al., 2000; Creel et al., 2002) have investigated the effect of recreational activities on such species. Creel et al. (2002) found that the levels of glucocorticoids (indicating physiological stress) in wolves were substantially higher at locations and times with more intense touristic activity. According to a study run by Linnell et al. (2001), the Eurasian lynx is the only species of large carnivore in Europe for which a statistically significant correlation between human population density and historical extinction since the early 1800s has been observed. However, this was probably linked to a greater sensitivity to human influence on their ungulate prey rather than to disproportionate human persecution (Linnell et al., 2001). With regard to present coexistence with people, at a European scale no relationship was found between human density and the status of lynx populations (Linnell et al., 2001). At a population scale, lynx proved to chose habitats with a medium level of human presence (Basille et al., 2009), probably as a consequence of the habitat choice of their main prey, the roe deer (Capreolus capreolus), which benefit to a large extent from current human land use practices (Mysterud, 1999). On the other hand, the only study conducted at a finer scale (Sunde et al., 1998) indicated that resting lynxes avoided humans, showing different responses to disturbance depending on the type of habitat. Therefore, the mechanisms of avoidance for this species may work at a finer scale, both spatially and temporally, as was found for wolves (Theuerkauf et al., 2003). Furthermore, different human activities may have different effects (Bunnefeld et al., 2006) and the sensitivity to disturbance may vary while animals are performing different activities. Bunnefeld et al. (2006) analyzed the places chosen by lynx when hunting or resting in an area characterized by a low human density. There, lynx did not avoid the surroundings of agricultural fields and roads, where the main activity was in the form of wheeled vehicles, but they avoided the surrounding of permanently occupied houses, that were the most consistent source of human activity in the area (Bunnefeld et al., 2006), i.e. likely the places with the highest probability of encountering humans per se. No study to date has focused on the potential effect of intense touristic activity on lynx behavior; nonetheless, at least during the main touristic season, proximity to tourist trails may also be linked to a high probability of encountering humans. The present study aimed to investigate the potential effect of tourism and traffic on two aspects of the Eurasian lynx’s behavior: the hunting of an ungulate prey and the choice to move and find a suitable daytime resting site on the days when a large prey was consumed. Firstly, we hypothesized that at night, when human activity is low, lynx hunting behavior may not be influenced by the presence of roads or trails and therefore lynx may kill ungulate prey independently of the proximity to such structures. Secondly, we focused on the mean distance moved by the lynx from a kill site to the corresponding daytime resting

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sites (hereafter: distance 'kill site to daytime resting sites'). As large prey provide the lynx with food for several nights (Jobin et al., 2000) and the pattern of lynx activity was found to be shaped mainly by searching for and consuming large prey (Schmidt, 1999; Jędrzejewski et al., 2002), the need to move strongly decreases while the prey is eaten (Schmidt, 1999). However, during the day lynx may be led to move to safe shelters due to human activity in the surroundings of the kill site. Therefore we hypothesized that the distance 'kill site to daytime resting site' may increase when the kill site is located near a tourist trail or a road. Also, we assessed whether this distance was influenced by habitat features at the kill site (whether they give the lynx the possibility to hide or not; Sunde et al., 1998) and whether there were differences between male and female lynxes. Finally, as individual variations have been observed in the rhythms or levels of activity among lynxes of the same sex, age and status (Schmidt, 1999), we also checked for individual differences. Material and methods Study area This study was conducted in the Bohemian Forest, a forested mountain range in South–West Czech Republic, along the border with Germany (48° 55' – 49° 17' N, 13° 13' – 13° 47' E). This region encompasses the Šumava National Park (680 km²) and a surrounding wide belt of Protected Landscape Area (PLA, 990 km²), where the main human activity is tourism and, to a less extent, forest management outside the non–intervention zones of the National Park. The foothills around the PLA are characterized by a denser net of paved roads, several small human settlements and a stronger influence of forestry and agricultural activities. Touristic activity is marginal in the surroundings closest to the PLA (10 km belt), except for a few renowned localities close to the main towns. Also due to this region’s recent history, the mean human population density of the whole area is low: about 20 ind./km2 and only 1.9 ind./ km2 in the central parts (Wölfl et al., 2001; Mašková et al., 2003). The Eurasian lynx is the only large carnivore species currently living in the area (Koubek & Červený, 1996) and it is present with a reintroduced population for which the estimated count, in 2001 was fewer than 70 individuals (jointly estimating the Czech, German and Austrian sides Wölfl et al., 2001). More recent data, obtained from a camera trapping project in the two National Parks Šumava and Bavarian Forest, led to an estimated density of 1.19 lynxes/100 km² (Weingarth et al., 2011). Throughout the whole region, the most common carnivore is the red fox (Vulpes vulpes), which is present in relatively high numbers. The primary species of wild ungulates are red deer (Cervus elaphus), roe deer —which is the main prey of the lynx (Okarma et al., 1997)— and wild boar (Sus scrofa). Among the mentioned species, the red fox is probably the most common mammal scavenger on

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lynx kills (Jędrzejewska & Jędrzejewski, 1998; Selva et al., 2005; Helldin & Danielsson, 2007) although wild boar often feed on carcasses as well (Selva et al., 2005). Lynx data During winters 2009/2010 and 2010/2011, lynxes were live–trapped using box–traps set on traditional lynx paths or at a fresh kill. The individuals were immobilized with 'Narketan' (10% ketamin), measured, fitted with GPS/GSM–collars (from Vectronics Aerospace, Berlin) and set free at the same place where they were captured. In this way we GPS–collared five lynxes: two females without kittens (F1 and F2), two adult males (M1 and M2) and one young male (M3). Collars were programmed to take one GPS position at midday, when the lynx is supposed resting (Schmidt, 1999), and one GPS position at midnight, when the lynx is supposedly active, hunting or moving through the territory (Schmidt, 1999). In addition to this schedule, for one month each season of the year, the collars took two additional GPS positions per day, at dawn and twilight, in order to obtain the complete series of ungulate prey killed by each collared lynx. Finally, every second week each animal had a 16–hour 'intensive monitoring period', when collars took one GPS positions per hour (from 4.00 p.m. to 8.00 a.m. at the following day). We obtained GPS positions from collars via SMS and downloaded them to a portable GPS (Trimble Juno SB), which was used to search for potential prey in the field. The actual prey site was then saved in the portable GPS. On the basis of the GPS fixes, from April 2010 to August 2011 we found 140 killed ungulate prey. In 116 cases collars also successfully measured the corresponding distance to 'daytime resting sites' of the lynx, represented by the midday GPS positions between two consecutive nights spent at the kill. For each of these 116 killed ungulate prey (fig. 1), we calculated the mean distance between the kill site and its corresponding distance 'kill site to daytime resting sites'. Explanatory variables We took into consideration nine potential explanatory variables (table 1), accounting for information about habitat structure and human activity in the surroundings of the kill sites and about potential changes in human activity and lynx behavior throughout the year. To evaluate the possibility for the lynx to hide near the kill site we adapted the 'cover pole method' (Pierce et al., 2004) to our aims: for each killed ungulate we verified in the field we placed a 2m–high pole, divided into 10 colored segments, at the location where the kill was found and we recorded how many segments were hidden for more than 50% when observed at a distance of 20 m in each cardinal direction, at a height of 1 m (fig. 2). By calculating the mean value for the four cardinal directions, we obtained an 'index of the presence of hiding places' ('habitat concealment', table 1). As lynx are known to prefer areas with steep

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Human settlements (diferent sizes) Random location

Sumava PLA

Kill locations

10 km belt outside PLA

National border

Paved roads

Bavarian Forest NP

Tourist trails

Sumava NP

Forest areas

CZ D

N 0 1 2

8 km

Fig.1. Distribution of the kill locations (n=116) and random locations (n = 116) throughout the Sumava National Park (NP), Protected Landscape Area (PLA) and a 10 km belt around the PLA. Fig. 1. Distribución de las presas cazadas por los linces (kill locations, n = 116) y de las localizaciones elegidas al azar (random locations, n = 116) en el Parque Nacional de Sumava (NP), en el Área de Paisaje Protegido (PLA) y en un cinturón de 10 km alrededor del PLA.

slopes (Basille et al., 2008) we also calculated the slope ('slope', table 1) for each kill site using a GIS layer with a 15 x 15 m–resolution (source: Český Úřad Zeměměřický a Katastrální–ČUZK, Praha). To take into consideration the potential effect of traffic and tourism, using specific GIS layers (sources: Český Úřad Zeměměřický a Katastrální–ČUZK, Praha and Sprava NP a CHKO Šumava) we calculated the distance of each kill site to the closest paved road (used mainly by motorized vehicles– 'Roads', table 1) and the closest tourist trail (used mainly by hikers and bikers). The most important factor influencing an animal’s behavior is likely to be the level of human activity on a road and not the presence of the road itself (Theuerkauf et al., 2003). Unfortunately, as we had no information about the actual number of tourists and vehicles using a trail or a road throughout the day and throughout the year, we used three alternative ways to include this information: (1) we divided the year into three periods (main summer tourist season from July to September, main winter tourist season from January to February and the two

periods in between, pooled together), according to the data collected by NP rangers about attendance at the main touristic localities ('period', table 1); (2) on the basis of our personal knowledge of the area and of existing GIS layers, we distinguished the 'main tourist trails' (mostly located inside the NP and PLA and around the main surrounding towns) from the 'irregularly used tourist trails' and we calculated the distance to tourist trails including first only the main tourist trails ('TurtraMAIN', table 1), then all the tourist trails without any distinction ('TurtraALL', table 1); and (3) because of the differences in the main human activities among the NP, PLA and the surroundings of PLA, we recorded in which of the three sub–areas the kill site was located ('area', table 1). Finally, about potential changes in lynx behavior, by distinguishing the kills that were found in the period from mid–January to the end of March from the others, we took into consideration the potential variation in the lynx behavior during the mating season ('mating_season'; table 1) as a period of higher mobility (Jędrzejewski et al., 2002).

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Table 1. Explanatory variables used in the analysis (see further details in the text): a Square–root transformed; b Log transformed. Tabla 1. Variables explicativas utilizadas para los análisis (para mayor detalle, véase el texto): a Transformada por raíz cuadrada; b Transformada logarítmicamente. Variable name

Description

Fixed effects Habitat Index of the presence of hiding places at the kill site, calculated adapting concealment a the 'pole method' (Pierce et al., 2004) Values: 0 (totally open habitat), 10 (totally closed habitat) Slope b Slope (with 15 x 15 m resolution) TurtraMAIN b Distance from the kill site to the closest main tourist trail (in m) TurtraALL b Distance from the kill site to the closest tourist trail, without distinction (in m) Roads b Distance from the kill site to the closest paved road Area NP (the kill site was inside the National Park) PLA (the kill site was inside the Protected Landscape Area) OUT (the kill site was in the surroundings of the PLA) Period Y1 (main summer tourist season) Y2 (main winter tourist season) N (both periods between the two main seasons) Sex

F (female), M (male)

Mating_season Y (the prey was killed during the mating season) N (the prey was killed during the rest of the year) Random effect Individuals M1 (adult male 1), M2 (adult male 2), M3 (young, non territorial male), F1 (adult territorial female), F2 (young territorial female)

Statistical analysis Using 'Hawth’s Tools' (Beyer, 2004) extension for ArcGIS 9.2 (ESRI, 2009) we created a set of random locations (n = 116) inside the whole study area (delimited by the National Border, by the border of the 10 km belt around the PLA and by the super home range of the five GPS–collared lynxes, estimated using the Kernel estimator at 100%). For each random location we calculated the distance to the closest paved road and the closest tourist trail (without any distinction), as we did for the kill sites (see above).We used Generalized Least Square regression (GLS) to test if the killed prey were located significantly closer or farther to roads and tourist trails than the random locations. Gaussian spatial correlation structure was specified to take spatial autocorrelation of individual observations into account. By means of Linear Mixed Effect Models (hereafter LME, fitted using the nlme package (Pinheiro et al., 2012) we tested if the mean distance 'kill site to daytime resting sites' was influenced by the distance to nearest paved roads or tourist trails, by the habitat structure, by the sex of the lynx and whether it varied

between the different periods of the year (mating season; main touristic season) or between the different sub–areas (National Park, Protected Landscape Area, closest surroundings outside the PLA). The identity of a given individual was included as random intercept into LME to take statistical non–independence of our data into account. LMEs did not exhibit any sign of spatial autocorrelation (assessed based on semivariogram). We therefore did not explicitly specify information on geographic locations of individual observations in LMEs. To achieve normality of residuals and homogeneity of variance the response variable (distance 'kill site to daytime resting sites') was Box Cox transformed (λ = 0.3). Square–root transformation was further used in the case of habitat concealment and log transformation in the case of distance to nearest paved roads and tourist trails and in the case of the slope. Although these transformations improved fits of evaluated models, their effect on the significance of individual exploratory variables were negligible. Collinearity is unlikely to affect our interpretations as correlation between explanatory variables was low; Spearman’s r ranged from –0.0150 to 0.1599 with the exception of the correlation between the distance

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Observing point S: 0 pole segments are > 50% hidden

1 m

Observing point N: 8 pole segments are > 50% hidden

2 m cover pole, divided into 10 colored segments

20 m

Killed prey W Fig. 2. Adapted version of the 'cover pole method' (Pierce et al., 2004): we obtained a measure of habitat concealment for each of the four cardinal directions and we then calculated the mean value. Fig. 2. Versión adaptada del “método del poste” (Pierce et al., 2004): obtuvimos una medida de la ocultación del hábitat para cada uno de los cuatro puntos cardinales, y luego calculamos su valor medio.

to the nearest tourist trails and the distance to the nearest main tourist trails (r = 0.7329). These two variables were, however, never included in the same model simultaneously (see below). We used a backward stepwise procedure to select the best minimal adequate model (hereafter MAM), i.e. the most parsimonious with all the effects being significant (Crawley, 2007). The significance of a particular explanatory variable was based on the change in deviance between the model containing this term and the reduced model, assuming a c2 distribution of the deviance change (Crawley, 2007). All calculations were carried out with the statistical software R 2.14.1. (R Development Core Team, 2011). Results Kill sites tended to be closer to tourist trails than random points (mean distance to all tourist trails for kill locations = 343 m, for random locations = 470 m – GLS: Δ df = 1, Likelihood ratio = 3.159, p = 0.0755), while we found no difference in the distance to paved roads (mean distance to paved roads for kill locations = 893 m, for random locations = 865 m – GLS: Δ df = 1, Likelihood ratio = 0.020, p = 0.8849). For each of the five collared lynxes we calculated the minimum, maximum and mean distance kill site to daytime resting sites', which are summarized in table 2. The mean distance values ranged from 752 m for female F1 to 1746 m for male M2 (table 2) and differed significantly between individuals (Anova: F(4,110) = 2.471, p = 0.0488). Finally, we tested the effect of each of the explanatory variables on the distance 'kill site to daytime resting sites' and we found that such distance was negatively related to the 'index of the presence of

hiding places' (habitat concealment, fig. 3A) and to the distance to all tourist trails (fig. 3B). The effect of other explanatory variables (area, mating season, roads, period, sex and slope) was not significant (table 3). Approximate r2 for the MAM was 0.201. The distance to main tourist trails tended to correlate more strongly with the distance 'kill site to daytime resting sites' when included into the model instead of the distance to all tourist trails (slope = –2.9578 ± 1.1570, Δdf = 1, c2 = 6.496, p = 0.0108), yet significance of other explanatory variables remained unchanged when including the former instead of the latter variable into the model. Discussion Regarding our first hypothesis, we found that the distribution of the kill sites was actually independent of the proximity to paved roads, and kill sites were even closer to tourist trails than random locations (almost significantly, see above). In several studies it has been observed that at night both carnivores and ungulate used at least the smallest and less frequented gravel roads, probably to move quickly and save energy (Stener, unpubl. data cited in: Sunde et al., 1998; Creel et al., 2002; Dickson et al., 2005). Sunde et al. (1998) also concluded that the avoidance of human facilities by the lynx is likely linked to the presence of people rather than to the alteration of the habitat (i.e. the mere presence of roads, trails and houses). Therefore, considering that people use human facilities mainly during the day and lynx mainly hunt during night time (Schmidt, 1999; Bunnefeld et al., 2006), our results are in accordance with these previous findings. During the day, we found that the longest mean and maximum distance 'kill site to daytime resting

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Table 2. Distances 'kill site to daytime resting sites' walked by each individual (in m): mean (± standard error), minimum and maximum distance. Na. Number of kills found for which the corresponding GPS daily positions were available; MD. Mean distance (mean ± SE); MinD. Minimum distance; MaxD. Maximum distance. Tabla 2. Distancias “lugar de caza a zonas de descanso diurnas” andada por cada individuo (en m): distancia media (± error estándar), mínima y máxima: Na. Número de presas halladas para las que estaban disponibles las posiciones diarias por GPS correspondientes; MD. Distancia media (media ± EE); MinD. Distancia mínima; MaxD. Distancia máxima. Monitoring period

MD kill–rest

MinD

MaxD

March 2010–May 2011

752 ± 153

3,716

March 2010–August 2011

1115 ± 131

4,249

March 2010–July 2010

936 ± 264

107

2,172

February 2010–March 2011

1746 ± 337

101

4,835

January 2011–August 2011

820 ± 146

1,709

sites' were walked by adult male M2 (table 2), which is consistent with the general result that territorial males generally travel longer daily distances than females (Schmidt, 1999; Jędrzejewski et al., 2002), probably due to their much larger territories (Schmidt et al., 1997). Data from a previous lynx radiotelemetry study in the Bohemian Forest (Bufka et al., in prep.) confirm this general result: the mean yearly home range size proved to be 438 km2 for adult males and 278 km2 for adult females, while the mean daily movement distance (DMD, sensu Jedrzejewski et al., 2002) was 11.5 Km for males and 6.5 for females. In the case of the second adult male, M1, we found the longest minimum distance 'kill site to daytime resting sites', while the mean distance value calculated for this male might be underestimated, as the data about M1 were collected only outside the mating season (table 2), when territorial males showed a longer locomotory activity than during the rest of the year (Schmidt, 1999; Jędrzejewski et al., 2002). In the case of the young male M3, the distances 'kill site to daytime resting sites' (mean, minimum and maximum) were shorter than the ones walked by both adult males, which is partially in contrast with what found by Schmidt (1999) when comparing adult and subadult males. This may be due to the low amount of data available for M3 (table 2) or simply to individual differences between lynxes (Schmidt, 1999). Nonetheless, a possible explanation may also be that M3 seemed to be still a floater: his home range included a large part of the territory held by male M2 minimally for the last three years and he used the whole home range in a very irregular way, spending several consecutive weeks in a restricted area and then moving far elsewhere (Bufka & Belotti, unpubl. data).Therefore, the movements of M3 may not have been influenced by the need to patrol a territory, which proved to be among the strongest motivations for movement in the case of males (Jędrzejewski et al., 2002). Both females, F1

and F2, walked very short minimum distances 'kill site to daytime resting sites' and female F1 walked also the shortest mean distance. This is consistent with findings by Jędrzejewski et al. (2002), that females in general walked shorter daily distances than territorial males, and findings by Schmidt (1999), that individual differences among lynxes of the same sex and status may play a role. Although previous studies (Schmidt, 1999; Jędrzejewski et al., 2002) found different daily activity and movement patterns for lynxes of different sex, age and status, with the data available for our study (five collared lynxes) we could only test for differences between male and female lynxes. We found no significant differences, although this could be due to the limited number of GPS–collared individuals. In general, lynxes seemed to react to the presence of tourist trails, where people mainly move on foot, by bike, or by cross country skiing: they walked farther during the day when the prey was located closer to a tourist trail and this negative correlation tended to be stronger when we considered only the main trails. Although the possible reaction of lynx to touristic activity has not been taken into consideration previously, this may be considered consistent with findings by Bunnefeld et al. (2006) that showed that lynxes in less urbanized areas avoided the surroundings of occupied houses much more than the surroundings of roads, probably because houses were associated with a higher probability of encountering people on foot. Schmidt (1999) also found that human disturbance may have an effect on the behavior of females with kittens: they observed that, during daylight hours, the level of activity was lower in females living in an area where human activity was high than in females occupying low density urbanized areas. Although there is also evidence that reactions to disturbance may vary for lynxes of different sex, age and status (Bunnefeld et al., 2006), as mentioned above we could not test for such differences because we had data on

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A Distance (kill site–daily resting sites)

1.0

3.0

40 Distance (kill site–daily resting sites)

1.5 2.0 2.5 Habitat concealment

1.0 1.5 2.0 2.5 Distance to tourist trails

3.0

Fig. 3. Relationship between the distance 'kill site–daily resting site' (Box–Cox transformed, λ = 0.3) and: habitat concealment (square–root transformed, A); distance to the closest tourist trail (log transformed, B). Regression slope and 95% confidence intervals correspond to LME based predictions. Fig. 3. Relación entre la distancia “lugar de caza a zonas de descanso diurnas” (transformada Box–Cox, λ = 0,3) y: cripticismo del hábitat (transformada por raíz cuadrada, A); distancia del lugar del sacrificio al sendero turístico más cercano (transformada logarítmicamente, B). La pendiente de la regresión y los intervalos de confianza del 95% corresponden a las predicciones basadas en LME.

too few GPS–collared lynxes. Also, because we could not count on accurate data about the real amount of people using a trail, we could not precisely determine whether the intensity of human activity plays a role. The fact that we found a stronger negative correlation when the distance 'kill site to closest tourist trail' was calculated only considering the regularly used tourist trails may indicate that tolerance to humans varies according to the intensity of human activity. Nonetheless, we also obtained a negative correlation

when we considered all tourist trails, and we found no difference between the mean distances 'kill site to daytime resting sites' walked by lynx during the two main tourist seasons and during the rest of the year. This may mean that the level of activity outside the main tourist seasons and even on the less used tourist trails is enough to have an effect on lynx behavior. Nonetheless, to clarify this aspect, more precise estimations of the spatial and temporal changes in the intensity of human activity are required.

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Table 3. Effect of the explanatory variables on the distance 'kill site to daytime resting sites' (Box–Cox transformed, λ = 0.3) based on the linear mixed effect model with normal distribution of errors. Backward elimination of non–significant terms was used to select the best minimal adequate model (MAM) with all its significant effects (see Methods for details). Models were compared using the likelihood ratio test. Significant factors included in the MAM are in boldface. Values of parameter estimates and their significances are statistically controlled for all effects included in MAM. Tabla 3. Efecto de las variables explicativas sobre la distancia “lugar de caza a zonas de descanso diurnas” (transformada Box–Cox, λ = 0,3) basado en un modelo de efectos mixtos lineal con una distribución normal de los errores. Se utilizó la eliminación posterior de los términos no significativos para seleccionar el mejor modelo adecuado mínimo (MAM), siendo todos sus efectos significativos (véase Métodos para más detalles). Se compararon los modelos utilizando el test de razón de verosimilitud. Los factores significativos incluidos en el MAM están en negrita. Los valores de las estimas de los parámetros y de sus significancias están controlados estadísticamente para todos los efectos incluidos en el MAM.

Estimate

± S.E.

Δdf

Likelihood ratio

(Intercept)

37.375

4.441

25.081

< 0.0001

TurtraALL_log

–3.163

1.481

4.547

0.033

Habitat concealment_rad2

–3.756

1.238

9.010

0.003

Roads_log

–1.045

1.645

0.415

0.520

Sex (females vs. males)

2.512

2.234

1.158

0.282

Mating_season (N vs.Y)

2.423

1.723

2.017

0.156

2.179

0.336

Period (N vs. Y1)

0.383

1.866

Period (N vs. Y2)

2.749

1.904

Area (CHKO vs. NP)

–0.554

1.805

Area (CHKO vs. OUT)

–1.890

2.091

Slope_LOG

–1.572

0.887

2 1

0.826 3.179

0.662 0.075

We found no effect of the proximity to paved roads mainly used by motorized vehicles. This may be explained by the fact that most human activity associated with paved roads is in the form of vehicles, which may not be perceived by animals as being as risky as humans per se (Andersen et al., 1996). Main paved roads (and highways, which were absent in our study area) are known to have a negative effect on dispersal and on connectivity among a species’ populations (Schadt et al., 2002), and this can be a serious problem for long–term conservation of large carnivores such as the Eurasian lynx (Breitenmoser et al., 2000). Nonetheless, paved roads are probably not perceived as a source of disturbance by animals performing activities such as feeding or resting (Bunnefeld et al., 2006). Concerning the habitat structure, we found that the presence of habitat features that are linked to a higher level of horizontal cover correlated negatively with the mean distance 'kill site to daytime resting sites' walked by the lynx. This is consistent with findings by Sunde et al. (1998) who reported that lynxes are able to tolerate the presence of people even at short distances if the habitat structure offers them sufficient cover. We found a marginally insignificant negative correlation between the slope and the distance 'kill site to daytime resting sites'. Consistently with this result,

Sunde et al. (1998) found that habitat inclination had no influence on the tolerance distance of lynx, and they concluded that the general preference for steep resting sites may be a result of preference for uncultivated forest stands, which are likely most abundant in steeper and less accessible portion of a forest. Indeed, lynx showed a stronger preference for steep slope in densely urbanized regions, where such features may be linked to a low level of human activity (Basille et al., 2008), than in areas with a low level of urbanization (Basille et al., 2009). In summary, our study indicates that even human activities that do not directly aim to damage wildlife (i.e. tourism) can influence lynx behavior. For the lynx, the choice to move longer distances to find a suitable resting site also on days when a large prey is available may have a cost in terms of 'energy expenditure through movements' and, in general, the choice to remain near the prey during the day may have the side effect of keeping scavengers away. In fact, few observations from camera trapping in our study area indicated that fox used lynx kills mainly after adult lynxes had left the site (Bufka et al., unpub. data). This is in accordance with a certain degree of temporal or spatial avoidance of lynx by foxes found in a few studies (discussed in Helldin et al., 2006).

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As the density of scavengers (red fox, wild boar) in the study area is rather high, choosing to remain near the prey may be an advantage, especially in the case of red deer prey and especially in winter as meat remains fresh for longer due to the lower temperatures. However, to determine whether recreational activity can actually have a negative effect on the fitness of the lynx, we believe that the next essential step is to test whether such activity can negatively influence the time lynx spend at the prey (how many times they come back to feed at the same prey and how long they stay at the prey during one night), as this is likely the key factor directly influencing animals’ fitness (Hik, 1995; reviewed in Frid & Dill, 2002). Using the lynx GPS positions from the 16–hour 'intensive monitoring periods' it is possible to investigate this aspect; nonetheless, the amount of confirmed ungulate prey that was found during such periods have been insufficient to allow this further step. Within our ongoing project, we aim to collect more data to answer this question. We also aim to compare the situation on both sides of the border between the Czech Republic and Germany, as lynx and ungulate populations often occupy trans–boundary territories and the two areas are part of the same mountain range, but show several differences (e.g. different human population density, primary landscape, landscape use, intensity of tourism). Over the last two decades, the tourist attractiveness of the Bohemian Forest has increased greatly and the flow of people on the road network and throughout the National border has intensified. This trend is likely to continue in coming years and there is currently pressure to restore several almost unused roads and trails. Therefore, we believe that the effects of human activity should be better studied in different species as they likely show different levels of tolerance to such activities (Frid & Dill, 2002; Taylor & Knight, 2003). Although more precise information about the amount of people using the different tourist trails would be required to determine such levels of tolerance, this study shows that this aspect of lynx coexistence with humans may also deserve more attention. In order to understand lynx responses to recreation further research is surely needed. Such information may help managers setting up recreation plans aiming to minimize the impact of human presence, a fundamental issue to achieve the best compromise between economic development and species conservation. Acknowledgements This project is financially supported by EU Program Interreg IV (Objective 3 Czech Republic – the Independent State of Bavaria) and by IGA CZU 2011 4315013123113. J. Kreisinger′s contribution was supported by institutional resources from the Ministry of Education, Youth and Sports of the Czech Republic for the support of science and research. We would like to thank O. Vojtěch, J. Mokrý, H. Burghart and M. Gahbauer, for their help in the field work. We also thank K. Mayer for cooperation during prey searching and useful comments to improve the manuscript.

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Influence of new irrigated croplands on wild boar (Sus scrofa) road kills in NW Spain V. J. Colino–Rabanal, J. Bosch, Mª J. Muñoz & S. J. Peris

Colino–Rabanal, V. J., Bosch, J., Muñoz, Mª J. & Peris, S. J., 2012. Influence of new irrigated croplands on wild boar (Sus scrofa) road kills in NW Spain. Animal Biodiversity and Conservation, 35.2: 247–252. Abstract Influence of new irrigated croplands on wild boar (Sus scrofa) road kills in NW Spain.— In recent decades, wild boar populations have increased both in number and distribution. This rise is partly related to the increase in cropland devoted to maize (Zea mays) cultivation, as wild boar find food and refuge in these areas. This population expansion has led to an increase in the number of wild boar vehicle collisions (WBVCs). The goal of the present study was to evaluate a set of spatio–temporal factors that influence WBVCs related to maize crops on the Northern Spanish Plateau (the region of Castile and Leon). We compared the maize pattern with the factors related to total WBVC numbers. We observed that whereas the total occurrence of WBVCs usually increased with forest cover and speed and traffic volumes, maize areas were one of the main explanatory variables in plateau models. To avoid collisions in these areas in future, a number of mitigation measures are outlined. Key words: Maize, Wild boar, Road–kill, Northern Spanish plateau, Population increase, Shelter areas. Resumen Influencia de los nuevos cultivos en regadío sobre las colisiones de vehículos con jabalí (Sus scrofa) en el Noroeste de España.— Las poblaciones de jabalí han aumentado en número y distribución en las últimas décadas en España. En dicho aumento, interviene la mayor extensión de los cultivos de maíz (Zea mays); donde la especie encuentra refugio y alimento. La expansión de la especie, provoca un aumento de colisiones de jabalí con vehículos (WBVC). El presente trabajo trata de la distribución espacio–temporal del WBVC en la meseta norte española. Normalmente, el WBVC se relaciona con la velocidad, el volumen de tráfico y/o cubierta forestal. Sin embargo, el maizal resulta la variable más relacionada en atropellos con jabalí en la meseta norte. Se dan algunas recomendaciones para evitar futuras colisiones en esta área. Palabras clave: Maíz, Jabalí, Atropellos en la carretera, Meseta Norte española, Incremento población, Refugios. Received: 19 XII 11; Conditional acceptance: 10 II 12; Final acceptance: 9 VI 12 V. J. Colino–Rabanal & S. J. Peris, Dept. of Zoology, Fac. of Biology, Univ. of Salamanca, 37071 Salamanca, España (Spain).– J. Bosch & Mª J. Muñoz, Animal Health Research Center, CISA–INIA, 28130 Valdeolmos, Madrid, España (Spain). Corresponding author: S. J. Peris. E–mail: peris@usal.es

ISSN: 1578–665X

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Introduction The wild boar Sus scrofa has an excellent capacity for adaptation and occupies an extremely wide range of habitat types. These features, together with its high reproductive rate and low hunting and predator pressure, explain the recent widespread increase in both the populations and their geographical range (Saez–Royuela & Telleria, 1986). They are currently found in more than two thirds of the Iberian peninsula and populations seem to be increasing in parallel with the increase in new irrigated land (Herrero, 2003). Such newly irrigated areas now make up more than 21.4% of Iberian croplands, and they are mainly devoted to maize (Zea mays), a crop which has increased by 10% in the past 15 years and now covers 349.163 ha (2008 data). Wild boars use these croplands as feeding and resting areas for up to 8 months of the year —from June to March, depending on prevailing agricultural practices— especially if other nearby resources are scarce. The population increase causes important economic losses and damage to croplands (Calenge et al., 2004; Herrero et al., 2006; Schley et al., 2008). Moreover, it has led to a growing number of wild boar vehicle collisions (WBVCs). In many parts of Europe, animal–vehicle collisions are an increasing traffic safety problem (Groot Bruinderink & Hazebroek, 1996). Together with roe deer, Capreolus capreolus, in many parts of Mediterranean countries wild boar is the main species involved in vehicle accidents (Colino–Rabanal, 2011). WBVCs tend to be concentrated in distribution and tend to occur close to feeding or resting sites on road segments with certain characteristics (Primi et al., 2010). They are also concentrated in time, with higher collision rates at night, during the rutting season, and during litter dispersion (Peris et al., 2005). WBVC distribution is related to the sex ratio, and the percentage of juveniles and adults is representative of the population structure (Groot Bruinderink & Hazebroek, 1996). To reduce WBVCs, several mitigation measures have been proposed (Huijser et al., 2009). Aiming to identify road segments where mitigation measures might be implemented, previous research has focused on identifying the variables involved in WBVCs (Primi et al., 2010). In such studies, WBVCs have been considered as a whole and the possibility of different WBVC patterns has not been taken into account. Different patterns , are to be expected owing to the wide range of habitats the species is able to occupy. The most suitable mitigation measures may therefore vary depending on these patterns. We hypothesized that the spatio–temporal variables involved in WBVCs in relation to maize croplands may differ from those involved in WBVC located in other areas. Knowledge of the specific factors involved in this particular pattern would enable road planners and environmental managers to improve mitigation measures based on the specific conditions of WBVCs in maize areas. In a previous study, we defined several WBVC patterns in the region of Castile and Leon (NW Iberian peninsula, 94,223 km2, 26.57 inhabit./km2) using WBVC traffic reports and neural networks (Colino–Rabanal, 2011). One pattern was

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related to maize crops on the plateau (at least 10% of all WBVCs within the region were clearly related to maize crops). The goal of this study was thus to describe the spatio–temporal distribution of WBVCs related to irrigated maize croplands and to compare it with the results for total WBVCs. Material and methods We worked with WBVC traffic reports from 2002 to 2008 made by the Spanish traffic safety authorities (Guardia Civil). Each report included at least the WBVC location (the name of the road and the kilometer point) and the time it took place (date and hour). Accuracy of the WBVC location is ± 50 m (data recorded in hectometers). This accuracy helps to eliminate misleading results (Gunson et al., 2009). We used a dataset from a geographical information system. The software used was ArcGis 9.0. The WBVC location sites for both total numbers and maize patterns were modeled using stepwise logistic regressions. We compared the WBVC spatial distribution (n = 2,347 WBVC) with a random–point distribution with the same number of control points. Before modeling, we checked for multicolinearity between variables. If the correlation coefficient was higher than 0.70, we removed one of the variables. We included variables related to traffic, human disturbances, and landscape uses. The variables included in the models are explained in table 1. The traffic parameters came from traffic intensity maps provided by the Regional Government of Castile and Leon and the Spanish Ministry of Public Works. We obtained land use data from the Forest Map of Spain (scale = 1:50,000) drawn up by the Spanish Ministry of the Environment during 2002 and 2003. As a sample unit for landscape variables we used a circular surface with a 1,000–meter radius around the WBVC localities. For the maize crops, we used the proportion of area occupied in the municipality where the WBVC took place. These data, the number of hectares for each regional municipality, were provided by the Agricultural Department of the Regional Government of Castile and Leon. Figure 1 shows the spatial distribution of irrigated crops in Spain and the WBVC within the region of Castile and Leon. Results Recent trends and temporal patterns WBVCs have increased both on croplands and within the whole region (fig. 2). WBVCs were not equally distributed during the year but tended to concentrate during autumn, from October to January, for both total WBVC numbers (c2 = 55.67, d.f. = 11, p < 0.001) and maize patterns (c2 = 114.51, d.f. = 11, p < 0.001). In the case of WBVCs related to maize, this trend was even more pronounced for November and December (the percentage of WBVCs occurring in these months for maize pattern was higher than for the total) (fig. 3).

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Table 1. Traffic, environmental and anthropogenic variables measured regarding both wild boar vehicle collision (WBVC) locations and a random distribution of the same number of points as control sites. Tabla 1. Variables del tráfico, ambientales y antropogénicas medidas tanto en los puntos donde se han registrado colisiones de jabalí con vehículos (WBVC) como en aquellos que se corresponden con la distribución de puntos al azar. Variable

Variable explanation

Traffic volume

Volume of vehicles on the road per day (vehicle/day)

Slope

Average speed of vehicles (km/h)

Road density

Road length per unit surface area (km road/km2)

Railway density

Railroad length per unit surface area (km railway/km2)

Sinuosity

Relationship between real and Euclidean distance

Slope

(25–metre resolution)

Distance from water

Euclidean distance to nearest river or stream (m)

Anthropogenic distance

Euclidean distance to nearest anthropogenic structure (m)

Forest

Proportion of natural forest coverage (percentage)

Reforestation

Proportion of reforestation coverage (percentage)

Wooded pasture

Proportion of wooded pasture coverage (percentage)

Riparian forest

Proportion of riparian forest coverage (percentage)

Crops

Proportion of crop coverage (percentage)

Maize crops

Area occupied in the municipality (hectares)

Scrub

Proportion of scrub coverage (percentage)

Pasture

Proportion of pasture coverage (percentage)

Mosaic landscape

Proportion of mosaic landscape coverage (percentage)

Edge density

Density of ecotones (km/km2)

Landscape diversity

Diversity of land uses (Shannon index)

Likewise, regarding the time of the day WBVCs were concentrated during the first hours of the night in both patterns (from 19:00 to 23:00 h) (fig. 3). Spatial patterns WBVCs were uncommon in agricultural landscapes except around irrigated maize crops. Table 2 shows the results of the best model for this WBVC pattern. Apart from the maize area, other variables included in the models were traffic volume and edge density. The results were different from those obtained for the total number of WBVCs, where traffic variables and the proportion of forest and mosaic landscapes were seen to be the main explanatory factors. Discussion Wild boars show high densities in irrigated maize croplands in the surroundings of the main rivers on the Spanish Northern Plateau. In maize croplands,

wild boar populations have good access to food resources, with stable year–round food availability (Herrero, 2003). Although in recent years the number of hectares devoted to maize growing has tended to stabilize, the recent increase in this crop has led to serious increases in WBVCs. This WBVC pattern has been given little attention in previous research and the mitigation measures proposed for other areas might not be the appropriate for the region studied here. Huijser et al. (2007) published a good review on available mitigation measures. It would stand to reason that not all measures are equally appropriate for each specific situation. We observed that WBVC increasee more in maize–growing areas than in areas with forest cover and/or highest speed and traffic volumes; maize was the main explanatory variable in the models. Moreover, apart from the proportion of maize crops on the landscape, other factors not considered in the models may contribute to explaining the occurrence and seriousness of WBVCs. For example, croplands in close proximity to the roadside probably increase the risk of WBVCs because they

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N W

E S

120

240 km

Corine Agricultural areas Permanently irrigated land

Fig. 1. Spatial distribution of irrigated croplands within the Iberian peninsula wild boar vehicle collisions (WBVC) in the study area (Castile and Leon). Surfaces covered by irrigation lands are depicted in grey. Fig. 1. Distribución espacial de las zonas de regadío en la península ibérica y de las colisiones de jabalí con vehículos (WBVC) en el área de estudio (Castilla y León). En ambos, las zonas de regadío se representan en gris.

1,600

Number of collisions per year

1,400

1,200 y = 52.857x + 977.29 R2 = 0.3467

1,000 800 600 400 200 0

y = 13.571x + 214 R2 = 0.4192 2002

2003

2004 2005 Year Maize

2006

2007

2008

Total

Fig. 2. Comparison between the temporal evolution of total wild boar vehicle collisions (WBVC) and those related to maize croplands. Fig. 2. Comparación entre la evolución temporal del total de las colisiones de jabalí con vehículos (WBVC) y de las relacionadas con los cultivos de maíz.

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J Months 25 D F

Hours 22

2 3 4

20 15

5 0

17 S

8 15

9 14

J Jl

Fig. 3. Temporal distribution of wild boar vehicle collisions (WBVC): by months and by the time of day. The black line refers to the maize pattern and the grey one to total WBVCs. Fig. 3. Distribución temporal de colisiones de jabalí con vehículos (WBVC): por meses y por hora del día. La línea negra indica cultivos de maiz y la línea gris el total de WBVCs.

limits drivers’ vision, and at the same time, they may attract more animals to areas close to roads. We acknowledge that the way in which we quantified the 'maize' variable, i.e., as the proportion of municipal area dedicated to maize cultivation could introduce some bias to the analysis. Previous research has usually measured the distance to a specific habitat or the proportion of habitat within a buffer of a certain radius. Moreover, crop rotation adds a temporal bias, which makes it more difficult to create precise maize maps. In these intensive agricultural areas farmers usually cultivate species in rotation, mainly maize and alfalfa.

We suggest that to reduce the number of collisions in the futre, apart from other more common mitigations such as road fencing, measures should be implemented to prevent sowing maize within a band of a certain width along roadsides in order to decrease the presence of the species and increase driver reaction time. The economic costs of this maize–free band could be calculated in terms of harvest loss; each 1 m of band–width, for example, amounts to a maximum of 1,200–1,500 kg of harvest loss per 100 m of road segment (with an average 4–6 plants/m2). At a cost of 12 €/1 Tm (data from September 2011, Spanish agriculture market),

Table 2. Variables included in the most parsimonious model for wild boar vehicle collisions (WBVC): A. Total WBVC data within the region. B. Maize crop patterns on the northern plateau. Tabla 2. Variables incluidas en el modelo más simple para las colisiones de jabalí con vehículos (WBVC): A. Total de WBVC en la región; B. Cultivos de maíz de la meseta norte.

A B SE Wald p–value Slope

–0.029 0.005 32.827

0.001

B B SE Wald p–value Traffic volume 0.765 0.190 16.126 p < 0.001

Traffic volume 0.001

0.001 174.810 p < 0.001

Maize

2.452 0.283 75.260 p < 0.001

Speed

0.045

0.006 66.021 p < 0.001

Edge density –1.204 0.382 9.915

Forest

0.322

0.032 100.955 p < 0.001

Mosaic

0.109

0,015 56.309 p < 0.001

0.002

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the reduction in cropland due to the band would reach 18–20 €/100 m road for each meter of band–width. This could be more affordable than other mitigation measures such as bridges, underpasses and so on. Cost–benefit analysis can be used to define the road segments where the installation of mitigation measures would be most profitable. In the present analysis, we compared the monetary costs and benefits of a certain mitigation measure, obtaining such benefits by combining the effectiveness of the measure in reducing the number of WBVCs and the costs associated with average WBVCs (Huijser et al., 2009). In this sense, farmers should be compensated for not planting in this band. In the region of Castile and Leon the costs associated with WBVCs are assumed by drivers if they do not obey the traffic code, by the hunting club if the animal comes from a private hunting preserve, or by the Regional Government if the WBVC is located within a natural park or if the risk of collision is not correctly signposted. In certain WBVC hotspots, either hunting clubs or the Regional Government will find it profitable to compensate farmers for not planting maize near the road rather than to cover r the economic damage derived from WBVCs. Future research should focus on identifying the best mitigation measures for collisions related to maize crops. References Calenge, C., Maillard, D., Fournier, P. & Fouque, C., 2004. Efficiency of spreading maize in the Garrigues to reduce wild boar (Sus scrofa) damage to Mediterranean vineyards. European Journal Wildlife Research, 50: 112–120. Colino–Rabanal, V. J., 2011. Contributions to the analysis of vertebrate road mortality. Ph. D. Thesis, Salamanca Univ. Groot Bruinderink, G. W. T. A. & Hazebroek, E., 1996. Ungulate traffic collisions in Europe. Conservation Biology, 10: 1059–1067. Gunson, K. E., Clevenger, A. P., Ford, A. T., Bis-

Colino–Rabanal et al.

sonette, J. A. & Hardy, A., 2009. A comparison of data sets varying in spatial accuracy used to predict the locations of wildlife–vehicle collisions. Environmental Management, 44: 268–277. Herrero, J., 2003. Adaptación funcional del jabalí Sus scrofa a un ecosistema forestal y a un sistema agrario intensivo en Aragón. Publicaciones del Consejo de Protección de la Naturaleza de Aragón, serie Investigación, Zaragoza. Herrero, J., García–Serrano, A., Couto, S., Ortuño, V. M. & García–González, R., 2006. Diet of wild boar Sus scrofa L. and crop damage in an intensive agroecosystem. European Journal Wildlife Research, 52: 245–250. Huijser, M. P., McGowen, P., Fuller, J., Hardy, A., Kociolek, A., Clevenger, A. P., Smith, D. & Ament, R., 2007. Wildlife–vehicle collision reduction study. Report to Congress. U.S. Department of Transportation, Federal Highway Administration, Washington D.C. Huijser, M. P., Duffield, J. W. Clevenger, A. P. Ament, R. J. & McGowen, P. T., 2009. Cost–benefit analyses of mitigation measures aimed at reducing collisions with large ungulates in the United States and Canada; a decision support tool. Ecology and Society, 14(2): 15. [online] URL: http://www.ecologyandsociety.org/vol14/iss2/art15/ Peris, S., Baquedano, R., Sanchez, A. & Pescador, M., 2005. Mortalidad del jabalí (Sus scrofa) en carreteras de la Provincia de Salamanca (NO de España): ¿influencia de su comportamiento social? Galemys, 17: 13–23. Primi, R., Pelorosso, R., Ripa, M. N. & Amici, A., 2010. A statistical GIS–based analysis of Wild boar (Sus scrofa) traffic collisions in a Mediterranean area. Italian Journal of Animal Science, 8: 649–651. Saez–Royuela, C. & Telleria, J. L., 1986. The increased population of the Wild Boar (Sus scrofa L.) in Europe. Mammal Review, 16: 97–101. Schley, L., Dufrêne, M., Krier, A. & Frantz, A. C., 2008. Patterns of crop damage by wild boar (Sus scrofa) in Luxembourg over a 10–year period. European Journal Wildlife Research, 54: 589–599.

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Game species monitoring using road–based distance sampling in association with thermal imagers: a covariate analysis K. Morelle, P. Bouché, F. Lehaire, V. Leeman & P. Lejeune

Morelle, K., Bouché, P., Lehaire, F., Leeman, V. & Lejeune, P., 2012. Game species monitoring using road– based distance sampling in association with thermal imagers: a covariate analysis. Animal Biodiversity and Conservation, 35.2: 253–265. Abstract Game species monitoring using road–based distance sampling in association with thermal imagers: a covariate analysis.— Monitoring of game species populations is necessary to adequately assess culling by hunters in areas where natural large predators are absent. However, game managers have to control several species and they often lack of an efficient and convenient survey design method. Monitoring several species at that same time over large areas could thus be cost– and time–effective. We tested the influence of several factors during monitoring of three common game species, (wild boar, roe deer and red fox, using road–based distance sampling in association with thermal imagers. This pilot survey based on 20 night counts in five contrasting sites studied the effect of several covariates (species, thermal imaging, observer, group size, and habitat type) on the detection probabilities (= dp). No differences were observed between thermal imagers (dpJENOPTIK: 0.186, dpFLIR: 0.193) and group sizes (dp1ind.: 0.243, dp2ind.: 0.259, dp> 2ind.: 0.223), but we found differences between observers (dpobs1: 0.207, dpobs2: 0.274, dpobs3: 0.159). Expected differences were also observed between species (dpwild boar: 0.22, dproe deer: 0.35, dpred fox: 0.32) and between habitat type (dpforest: 0.27, dpedge: 0.74, dpopen: 0.35). Our results show that the detectability of low cost thermal imaging equipment is similar to that of more expensive methods, highlighting new possibilities for the use of thermal imagery by game managers. Although adjustments should be made to the study design our findings suggest that large–scale multi–species monitoring could be an efficient method for common game species. Key words: Road–based distance sampling, Thermal imaging, Game species, Detectability. Resumen Monitorización de especies cinegéticas utilizando el muestreo a distancia con base en una carretera, en combinación con imágenes termográficas: un análisis de covariables.— La monitorización de las poblaciones de especies cinegéticas es necesaria para evaluar adecuadamente las capturas de los cazadores, en zonas que carecen de los grandes depredadores naturales. Sin embargo, los gestores de la caza deben controlar diversas especies y a menudo carecen de un método de control con un diseño conveniente. Por lo tanto, la monitorización de diversas especies al mismo tiempo en áreas muy grandes podría ser eficaz desde el punto de vista de los costes y del tiempo. Estudiamos la influencia de diversos factores durante la monitorización de tres especies cinegéticas comunes (el jabalí, el corzo y el zorro rojo) utilizando un muestreo a distancia desde la carretera, en asociación con imágenes termográficas. Este examen piloto basado en 20 recuentos nocturnos en cinco lugares contrastantes estudió el efecto de diversas covariables (especie, termografía, observador, tamaño del grupo y tipo de hábitat) sobre las probabilidades de detección (dp). No se hallaron diferencias entre las imágenes termográficas (dpJENOPTIK: 0,186, dpFLIR: 0,193) y el tamaño de los grupos (dp1ind.: 0,243, dp2ind. : 0,259, dp> 2ind.: 0,223), pero sí entre los observadores (dpobs1: 0,207, dpobs2: 0,274, dpobs3: 0,159). También se observaron diferencias esperadas entre las especies (dpwild boar: 0,22, dproe deer: 0,35, dpred fox: 0,32) y entre los tipos de hábitat (dpforest: 0,27, dpedge: 0,74, dpopen: 0,35). Nuestros resultados demuestran que la detectabilidad de los equipos de termografía de bajo coste es similar a la de otros métodos caros, destacando nuevas posibilidades del uso de la termografía para los gestores de la caza. Aunque deberían realizarse ajustes en el diseño del estudio, nuestros hallazgos sugieren que la monitorización de múltiples especies a gran escala podría ser un método eficaz para las especies cinegéticas comunes. ISSN: 1578–665X

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Palabras clave: Muestreo a distancia desde carretera, Termografía, Especies cinegéticas, Especies de caza, Detectabilidad. Rebut: 19 XII 11; Conditional acceptance: 27 II 12; Final acceptance: 3 IV 12 Kevin Morelle, Philippe Bouché, François Lehaire, Vincent Leeman & Philippe Lejeune, Unit of Forest and Nature Management, Univ. of Liege (Gembloux Agro–Bio Tech), Passage des déportés 2, 5030 Gembloux, Belgium. E–mail: kevin.morelle@ulg.ac.be

Animal Biodiversity and Conservation 35.2 (2012)

Introduction Increases in large game populations are reported from all over Europe (Saez–Royuela & Telleria, 1986; Jȩdrzejewska et al., 1997; Panek & Bresiński, 2002). Due to the extinction of large predators in most western areas of Europe, human control of game species through hunting is required to manage their populations (Apollonio et al., 2010). To control these populations efficiently it is important to establish sampling and counting methods that accurately estimate population size on which shooting numbers can be based. In this context, multiple–species monitoring over larger areas offers a practical means to estimate changes in regional abundance of targeted species (Manley et al., 2004). Multiple–species monitoring is often used for conservation management purposes (Regan et al., 2008; Manley et al., 2004), but is rarely applied in the field of wildlife game management. Manley et al. (2004) demonstrated that coordinated multiple–species monitoring is a robust alternative to intensive single species surveys and can be time and cost–effective for local managers. The challenge while working at a regional or landscape scale is to build up a robust survey design that takes into account variability among habitats and seasons (Jones, 2011). Despite these limitations, if the survey design is approriate, large–scale multi–species monitoring could be an interesting tool to help local stakeholders manage populations of the main, large–game species. Game species can be counted in the field using various methods. Spotlight (Heydon et al., 2000) is the most popular direct method while pellet count (Acevedo et al., 2010; Heydon et al., 2000) appears to be the most popular indirect method. Nowadays, distance sampling (Thomas et al., 2010), that takes into account variation in visibility (assumed to decrease with perpendicular distance to the surveyed transect), is largely included in these two counting methods (Ruette et al., 2003; Marques et al., 2001) because it allows better precision in estimating animal densities. As this methodology is well–established and widely used, for further details we recommend reading the most recent paper of the Distance development team (Thomas et al., 2010). Modelling the detection probabilities only in relation to the distance to the transect can be limiting when the surveyed transect crosses different habitats (Parrott et al., 2011) or when the sampling design involves monitoring different species with different morphological characteristics (Parrott et al., 2011; Barbraud & Thiebot, 2009). Variations in detectability may indeed lead to upward or downward estimation biases (Ramsey & Harrison, 2004). Survey conditions (Bozec et al., 2011) as well as observers (Pagano & Arnold, 2009; Marini et al., 2009) can also have an impact on the detection function and should also be included in any probability detection model. Recent work by Marques et al. (2007) demonstrates the importance of including covariates into the detection function modelling process to increase the model’s precision, although pooling data by relevant covariates

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can also be an effective strategy to deal with variability between covariates, e.g. grouping data by species (Focardi et al., 2002) or by habitat types (Acevedo et al., 2008). In spite of the role covariates might play in unbiased estimates of population density (Kéry & Schmid, 2004), few studies really focus on how they impact the detection function. Application of distance sampling using roads as transects implies limitations and benefits. On one hand, using roads to count has been shown to potentially affect animal behaviour (Shanley & Pyare, 2011; Rost & Bailey, 1979; Coulon et al., 2008) and distribution (Venturato et al., 2010; Roedenbeck & Voser, 2008; Erxleben et al., 2010). It can thus violate the assumption of random animal distribution around the transect required by distance sampling; for this reason the use of roads as transect is often considered a convenience sampling approach (Anderson, 2001). On the other hand, using vehicle on roads provides a means to cover large areas in a short time (Butler et al., 2007; Ward et al., 2004; Heydon et al., 2000; Gill et al., 1997) and is known to cause less disturbances to animals than walked transect (Heydon et al., 2000; Marini et al., 2009). Recently, use of thermal imaging to survey wildlife has gained in popularity (Hemami et al., 2007; Focardi et al., 2001; Gill et al., 1997). First used from the air (Garner et al., 1995), the method has been adapted for ground counts and has proved to be effective for detecting small (European hares Lepus europaeus Pallas 1778) to large (white–tailed deer Odocoileus virginianus Zimmermann 1780, muntjac Muntiacus reevesi Rafinesque 1815, roe deer Capreolus capreolus Linné 1758, red deer Cervus elaphus Linné 1758, wild boar Sus scrofa Linné 1758) game species (Focardi et al., 2001; Collier et al., 2007; Gill et al., 1997; Hemami et al., 2007). The advantages of thermal imagery are the ability to detect animals at night when they are more active due to less human disturbance (Cahill et al., 2003; Gottardi et al., 2010; Doncaster & Macdonald, 1997; Keuling et al., 2008; Kavanau, 1971; Barrio et al., 2010; Boitani et al., 1994). Detecting animals in evenly dense cover is also straightforward with thermal imaging. Compared to other techniques (e.g. spotlight counts), imaging techniques also cause less disruption to animal behaviour during counting (Fournier et al., 1995; Ward et al., 2004; Gill et al., 1997). Despite these advantages, wide use of thermal imagery is limited because of the high cost of this equipment compared to other techniques (Focardi et al., 2001). The objective of this paper was to evaluate the potential sources of inaccuracy and bias during a multi–species (wild boar, roe deer and red fox Vulpes vulpes Say 1823) monitoring survey over large areas (± 5,000 ha). We compared the detectability of two thermal imagers that differed in relation to their spatial resolution, visual comfort and price. We also tested for differences in detectability between species (expecting differences regarding their morphology), cluster size (expecting larger group to be more detectable), habitat (expecting higher detectability in open areas), observers (no differences expected) and time of the night (no a priori expectation).

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a b

c e

Surveyed sites Condroz

Fig. 1. Localisation of Belgium (left) and the five selected sites within the Condroz eco–region (right). Fig. 1. Localización de Bélgica (izquierda) y de los cinco lugares seleccionados en la región protegida de Condroz (derecha).

Material and methods Study area We conducted our study at five sites in Condroz, a natural region in Belgium situated between the Ardennes and the Meuse River (fig. 1). Condroz is a mosaic of woods and farmland, with 55% of its area consisting of grassland and crops (maize, cereals, beetroot and oilseed rape). The forest is patchily distributed and covers 24.5% of the total area. It lies at between 50 and 350 m above sea level. It has a temperate sub–oceanic climate with a mean annual temperature of 8°C and a mean monthly temperature varying between 2 to 16°C. The mean annual rainfall is 900 mm, and the mean annual duration of snow cover is over 25 days. The study sites covered an extension of between 4,600 ha (fig. 2, site a) and 6,400 ha (fig. 2, site d) with varying forest cover (ranging between 14 and 46%) and were limited either by natural or man–made barriers (highways, large rivers). Impermeability of these barriers is not assured, but for the purpose of this study we assumed that during fieldwork (from 28 II 11 to 15 IV 11) the population of the three studied species remained constant. Species monitored In these evenly wooded landscapes the most commonly hunted species are wild boar, roe deer and red fox. No culling strategies currently exist for these three species. Red deer are rarely present in the area and constitute a marginal hunted species in Condroz. Road transect sampling Within each study site we randomly selected survey transects among potential road segment (= road

network subdivided into 200 m long segment) candidates (fig. 3A) using the following four criteria: (i) distance to forest less than 300 m, (ii) distance to human settlements more than 100 m, (iii) paved, and (iv) low traffic roads. According to these criteria we then pooled each road segment into a suitable or unsuitable group (fig. 3B). For each site we defined a survey route using random selection among the suitable segments. Among the potential survey route candidates we selected the one that maximized the ratio [total length of suitable road segments/total length of survey route] (fig. 3C). If selected survey segments were parallel, we checked for a minimum distance of 500 m between them (Marini et al., 2009; Gill et al., 1997) to avoid potential double counting of animals. When not possible, a new route survey was generated until this condition was fulfilled. Selected survey routes ranged between 40.1 and 63.9 km. These lengths were designed to cover the habitat availability, to allow completion of the count within the lifetime of the thermal imaging battery (approx. 4–5 hours) and to ensure the constant attention of observers and vehicle driver during the count. Road density (mean = 6.3 km/km²) in the study areas was assumed to be sufficient to cover the different habitat conditions of each surveyed site. A mean of 0.9 km/ km² was surveyed across study sites, representing a mean sample rate of 14.4% of all (suitable and unsuitable) road segments. A four–wheel vehicle driven at low speed (10–15 km/h) was used to survey the designed road transects. One driver and two observers, one on each side of the vehicle, equipped with a hand–held thermal imager, were required to conduct the night count. Two different thermal imaging devices were used to detect animals, a FLIR ThermaCAM™ HS–324 with a resolution of 320×24 pixels, and a JENOPTIK VarioCAM™ with a resolution of 640×480 pixels. The main differences between the two

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Human Fields Water Forest Grassland 0 0.5 1

2 km

Fig. 2. Habitat composition of the five sites selected for the road–based distance monitoring. Selection was based on their homogeneous size, their gradient in forest cover (A = 46%, B = 37%, C = 25%, D = 18%, E = 14%), and their being limited by either natural or man–made barriers. Fig. 2. Composición del hábitat de los cinco lugares seleccionados para la monitorización desde la carretera. La selección se basó en su tamaño homogéneo, su gradiente de cubierta boscosa (A = 46%, B = 37%, C = 25%, D = 18%, E = 14%), y por estar limitadas por barreras naturales o hechas por el hombre.

cameras is the price (4585 € for the FLIR vs. around 32,000 € for the JENOPTIK), the weight (660 g for the FLIR vs. 1500 g for the JENOPTIK), the frame rate (8.3 Hz for the FLIR vs. 50 Hz for the JENOPTIK) and the visibility (one–eye viewfinder for the FLIR vs. 3.5” thin–film transistor (TFT) display for the JENOPTIK). Practically, these characteristics imply better visual comfort with the JENOPTIK camera with possibility of several colour ramps, but lower portability due to its weight. The FLIR camera is much cheaper and lighter but requires slower driving and more breaks for observers due to the lower resolution and eye fatigue using the viewfinder. Once an animal was detected, a spotlight combined with the laser rangefinder VECTOR IV Nite® by Vectronix were used to measure the bearing and distance between the animal and the road. We recorded the habitat context of a sighting (forest, open, edge) and also the fleeing response of the animal to spotlight use. Distinction between habitat contexts was based on animal (or cluster) position in relation to the forest. If an animal was in a 5 m–wide buffer in or outside the forest it was considered as being at the edge, while before or after this limit it was considered as being in open or forest habitat, respectively.

A GPSmap 62 receiver (Garmin™) was used to record the location of each sighting. Nocturnal road counts were conducted between 28 II 11 and 15 IV 11. This survey period was expected to be appropriate for surveying all three species because of the absence of vegetation in forest (favouring better visibility with thermal imagers) and the expected use of open areas (e.g. grasslands) at this time of the year (Baubet et al., 2004; Lucherini & Crema, 1994; Barancekova et al., 2010). To account for changes in animal activity throughout the night, four surveys were conducted in each area; two surveys took place at each site in the first half of the night (between 8 pm and 1 am) and two in the second half of the night (between 1 am and 7 am). For each particular site, the four counts were completed with a minimum of three days between each count. Between each count at a particular site, we alternated the starting point of the survey route (Marchandeau & Gaudin, 1994) and also the observers’ position in the vehicle to avoid potential bias. All surveys took place under similar weather conditions (dry conditions and temperatures ranging from 2 to 7°C). We did not therefore consider weather as a covariate in our analyses.

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Full road network 0 0.5 1 2 km

Total road network Suitable road segments

Random selection

Fig. 3. Stratified random selection process of the survey route: A. Full road network; B. Selection of suitable 200 m long segment (less than 300 m distance to forest, more than 100 m to human settlements, paved roads with little traffic; C. Random selection among potential segment candidates and final survey route. Fig. 3. Proceso de selección estratificada al azar de la ruta de seguimiento. A. Red de carreteras completa; B. Selección de segmentos de 200 m adecuados (a menos de 300 m de distancia al bosque, más de 100 m a poblaciones humanas, carreteras pavimentadas con poco tráfico); C. Selección al azar entre segmentos candidatos y ruta final del seguimiento.

Statistical analysis

Results

To estimate detection functions, we tested how animal detectability varied according to different levels of measured covariates: habitat type (factor levels: forest, edge and open), observer (obs1, obs2, obs3), time of the night (before/after midnight), camera type (JENOPTIK, FLIR), species (wild boar, roe deer, red fox), and cluster size (1, 2, > 2). The tested null hypothesis was the absence of difference between the detection probabilities of the levels of each of the considered covariates. Data were right truncated at 300 m. Perpendicular distance data and distance break classes were set to 50 m to smooth the detection function. Data collected during the 20 night counts did not allow us to study the effects of species and habitat type independently. Number of required sightings was indeed too low (< 30) to fit a detection function (Buckland et al., 2001). Accordingly, sightings data were pooled across sites and for each covariate analysis we pooled data independently of the species and the habitat type to ensure building detection function based on a sufficient number of sightings. For each covariate level we then tried to find the best detection function. We selected four potential candidate models: the half–normal, the uniform, the hazard rate and the exponential function. As our aim was to compare detectability and not to fit the best detection probability, no adjustment terms were added to these potential models. To fit these models we used the R package 'unmarked' (Fiske & Chandler, 2011). We used Akaike’s information criterion (AIC) to compare models and Mann–Whitney test to evaluate the differences (P < 0.05) in the detection probability between the covariate values tested.

Considering a sighting as any observation (an individual or a cluster of animals), we observed a total of 249 sightings: 42 were wild boar, 159 were roe deer, and 49 were red fox (table 1). European hares, rabbits (Oryctolagus cuniculus Linnaeus 1758) and occasionally badgers (Meles meles Linnaeus 1758) were also observed but not recorded. The mean perpendicular detection distance was 57.6 m (range = 0–230 m) for wild boar, 76.7 m (range = 0–435 m) for roe deer and 87.6 m (range = 5–266 m) for red fox (fig. 4). Detection distance among species varied significantly between wild boar and roe deer (Mann–Whitney test p–value = 0.0342) and between wild boar and red fox (Mann–Whitney test p–value = 0.0447). No difference between roe deer and red fox was observed. Detection distance was also affected by habitat (Mann–Whitney test p–value = 0.0007 for forest vs. open habitat and p–value = 0.0064 for forest vs. edge habitat), but edge and open habitat did not differ (Mann–Whitney test p–value = 0. 85). The mean cluster size was 1.84 for roe deer (range 1–5), 3.45 for wild boar (range 1–12), and 1.08 for red fox (range 1–2). Flight behaviour was observed in 56% of the sighting events for red fox, 31% for wild boar and 14% for roe deer. More than twice the number of sightings of roe deer was made after 1 am, while proportion of sightings for wild boar and red fox were similar before or after 1am (fig. 5). Covariate analysis We did not observe differences between the two thermal imagers (dpJENOPTIK: 0.186 ± 0.042 and dpFLIR 0.193 ± 0.043, fig. 6C) but we did find differences between observers (dpobs1: 0.207 ± 0.050, dpobs2: 0.274 ± 0.045, dpobs3: 0.159 ± 0.040, fig. 6D). For species

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Table 1. Transect characteristics (length of forest and open area surveyed), survey effort and total number of sightings and individuals. Tabla 1. Características del transecto (longitud del bosque y del área abierta estudiados), esfuerzo de seguimiento y número total de avistamientos de individuos. Site

Road transect (km)

Effort Visits (n) Survey (km)

Sightings (individuals)

Length

Forest

Open

48.4

29.4

193.6

5 (6)

42 (67)

24 (88)

64.4

19.2

45.2

257.6

3 (3)

68 (111)

2 (9)

63.9

9.9

255.6

18 (20)

18 (36)

6 (24)

40.1

8.3

31.8

160.4

10 (11)

15 (40)

7 (17)

44.6

4.2

40.4

178.4

12 (12)

15 (38)

3 (7)

Total

261.4

190.4

1045.6

49 (52)

158 (292)

42 (145)

covariates, the best model was the negative exponential for wild boar and red fox, and hazard–rate for roe deer (table 2). The probability of detection was significantly lower for wild boar (dpwild boar: 0.224 ± 0.053) than for roe deer (dproe deer: 0.351 ± 0.079) and red fox (dpred fox: 0.323 ± 0.046). Differences related to habitat types were also observed (fig. 6B). We did not observe any differences in detectability due to cluster size (fig. 6E) or time factors (fig. 6F) (table 2). Discussion Our results showed that thermal imagers with different characteristics can provide similar detectability. The higher spatial resolution of one camera over the other did not seem to benefit detectability, contrary to the expectations of Gill et al. (1997). This could be because it was used to detect animals by looking for 'hot spots' rather than identification per se (although in most case identification was possible when the car stopped). The difference in price for these two cameras and the comparable detectability shows the cost to use this technology can be greatly lowered, offering more possibilities for extensive use of this technology in the field, as suggested by Franzetti et al. (2011). However, although the FLIR imager is lighter than the JENOPTIK, observers found it more tiring to use because the viewfinder was more cumbersome than the TFT display of the JENOPTIK. Although it was not the purpose of our study, it is also important to mention that the FLIR imager did not allow us to classify detected animal by age or sex class as some imagers with higher resolution can do (Gill et al., 1997). In our study, thermal imagers were used for animal detection and for this purpose we argue that a lower resolution material can perform similarly. A camera with a medium resolution and a display rather than a viewfinder would be the best compromise. If different material is used in the field,

Fox

Roe deer Wild boar

we recommend preliminary tests be performed to confirm similar detectability. Our study confirmed that roe deer, wild boar and red foxes could be effectively detected during night count surveys with thermal imagers, but our detectability was somewhat lower than in other studies (Ward et al., 2004; Franzetti et al., 2011; Focardi et al., 2001). This could be related to the scale encompassed by our survey and the various habitat types crossed, as these other studies focused on one main habitat type (forest). The difference that we observed in mean detection distance between species might be due to animal behaviour. Red fox and roe deer were more often detected in open areas than wild boar. Our assumption that wild boar would use more grassland areas at this period of the year (Baubet et al., 2009) was not confirmed. The availability of earthworms for wild boar is higher under cold and rainy conditions (Baubet et al., 2003), and the dry weather conditions during our surveys could explain the low number of wild boar detected in open habitats. Observed differences in detection probability between habitat and species, although obvious and largely documented, emphasize the need to take the variations in cover and animal properties into account to estimate animal density with distance sampling. Differences in behaviour and shape between species can have an impact on detectability and therefore prevent inter–species pooling, although such pooling can sometimes be performed for other animal taxa (Oppel, 2006). We observed that the time of night seemed to have a potential impact on the number of sightings for roe deer, but not for wild boar or red foxes. In comparison, Heydon et al. (2000) found evidence of a time effect on fox sightings numbers, with more observations after midnight. These peaks of activity for a species can vary locally and should be carefully assessed before allocating survey efforts throughout the night. When the monitoring is done from the

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400

Distance (m)

300

200

100

Wild boar

Roe deer

Red fox

Fig. 4. Box plot of sighting distances for the three surveyed species. Fig. 4. Diagrama de cajas de las distancias de avistamiento de las tres especies estudiadas.

road, the greater pattern of animal activity may also be the consequence of less traffic between midnight and dawn (Vanhove & Ceuster, 2003). In previous studies the influence of observers on the detection function showed to impact density estimation (Ringvall et al., 2000). This influence is mostly due to systematic and random errors. In our study we observed differences between observers,

highlighting the need to train observers (by means of field trials) before starting the real survey so as to avoid this bias (Franzetti et al., 2011). In contrast with red fox, most roe deer and wild boar did not show flight behaviour when spotlighted. This confirms that spotlighting has a weak affect on roe deer behaviour during count (Ward et al., 2004; Smart et al., 2004; Gill et al., 1997) although several factors,

Proportion of sighting

1.0

Before 1 am After 1 am

0.8

0.6

0.4

0.2

Roe deer

Wild boar

Red fox

Fig. 5. Effects of time of the night on the number of sightings. Fig. 5. Efectos del momento de la noche sobre el nĂşmero de avistamientos.

Animal Biodiversity and Conservation 35.2 (2012)

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1.0

Roe deer (hazard–rate) Wild boar (exp. neg.) Red fox (exp. neg.)

0.8

Detection probability

1.0

0.6 0.4 0.2

0.8 0.6 0.4 0.2 0

0 0 50 100 150 200 250 300 Distance (m) D C 1.0 0.8

0.6 0.4 0.2

Detection probability

0.4 0.2

0 0

100 150 200 250 300 Distance (m)

100 150 200 250 Distance (m)

300

Obs. 1 (hazard–rate) Obs. 2 (exp. neg.) Obs. 3 (exp. neg.)

0.8 0.6 0.4 0.2

0 0 0 50 100 150 200 250 300 0 Distance (m) E F 1.0 1.0 1 indiv. (hazard–rate) 2 indiv. (exp. neg.) 0.8 0.8 > 2 indiv. (exp. neg.) 0.6

1.0

Jenoptik (exp. neg.) Flir (exp. neg.)

Detection probability

Forest (half–normal) Edge (exp. neg.) Open (exp. neg.)

100 150 200 250 Distance (m)

300

Before midnight (exp. neg.) After midnight (hazard–rate)

0.6 0.4 0.2 0 0

100 150 200 250 Distance (m)

300

Fig. 6. Detection probability function for (from upper left to lower right) species (A), habitat (B), thermal imager (C), observer (D), group size (E), and time of the night (F). Fig. 6. Función de la probabilidad de detección para (desde arriba a la izquierda hasta abajo a la derecha): especie (A), hábitat (B), técnica termográfica (C), observador (D), tamaño del grupo (E) y momento de la noche (F).

such as weather conditions and vegetation structure, have been shown to impact roe deer flight distances (De Boer et al., 2004). Wild boar have also showed low reaction to operators walking along transects (Marini

et al., 2009), and it has been observed that cars have less impact on animal behaviour than walkers when surveying owl (Manning & Kaler, 2011). We might expect the same conclusion in the case of large mam-

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Table 2. Detection probability values for the different covariates tested during the road–based transect survey. Detection probabilities (DP) with the same letter do not vary significantly (P < 0.05): 1 Akaike Information Criteria; 2 Effective half–strip widths: distance from the line at which the number of animal detected equals the number of animals missed. Tabla 2. Valores de la probabilidad de detección para las distintas covariables estudiadas durante el seguimiento en transectos con base en la carretera. Las probabilidades de detección (DP) con la misma letra no varían significativamente (P < 0,05): 1 Criterio de Información de Akaike; 2 Amplitudes efectivas de medio segmento: distancia de la línea en la que el número de animales detectados es igual al número de animales dejados pasar.

Covariates

Sample size (n)

Model

AIC 1

DP 2

ESWH (m)

Mean

Species Wild boar

Exp. neg.

90.59

56.1

0.224

0.053a

Roe deer

158

Hazard–rate

124.84

88.7

0.351

0.079b

Red fox

Exp neg.

58.14

94.6

0.313

0.046b

Habitat

Forest

125

Half–normal

135.07

60.9

0.271

0.092a

Edge

Exp. neg.

39.37

210.6

0.742

0.029b

Open

Exp. neg.

69.64

119.4

0.354

0.038c

Thermal imager

Jenoptik

144

Exp. neg.

129.49

55.8

0.186

0.042a

Flir

105

Exp. neg.

79.44

57.8

0.193

0.043a

Observer

Obs. 1

133

Hazard–rate

95.48

62.2

0.207

0.050a

Obs. 2

Exp. neg.

61.89

82.1

0.274

0.045a

Obs. 3

Exp. neg.

57.44

47.7

0.159

0.040b

Cluster size

138

Hazard–rate

103.48

72.9

0.243

0.057a

Exp. neg.

70.81

77.6

0.259

0.045a

> 2

Exp. neg.

96.47

67.0

0.223

0.044a

Period of night Before midnight

Exp. neg.

90.15

89.5

0.30

0.046a

After midnight

150

Hazard–rate

140.6

65.6

0.22

0.056a

mals accustomed to vehicle sounds. We were unable to find studies that compared motorized and walked transects and found effects on behaviour. Red fox fleeing behaviour has already been reported by other studies (Ruette et al., 2003). High hunting pressure in the study area on wild boar and red fox is likely to motivate flight when disturbed. It is important to note that most flight behaviour occurred after spotlighting and not at the detection time with the thermal imager. This suggests that using a thermal laser rangefinder to measure distances and bearing (instead of spotlight and daylight rangerfinder) may substantially decrease the flight response of animals during count.

Our study design did not allow us to collect sufficient data to estimate the density of the three game species at each site. As no other data on densities were available in our study sites, it remains difficult to know why our methodology did not succeed in collecting sufficient sightings for wild boar and red fox. Was it because the study design was too poor in terms of survey effort? Was it because of the elusive behaviour of these species? Or was it simply because their density was too low in the surveyed sites (Gill et al., 1997)? For further investigations, we suggest parallel studies should be conducted to collect data on abundance for the studied spe-

Animal Biodiversity and Conservation 35.2 (2012)

cies as this could help to make useful comparisons between methods. Moreover, the low number of wild boar sightings in more open landscape highlights the need for an adapted monitoring design for this species. Increasing the survey effort overall in forest habitat could, for example, favour the number of contacts with wild boar. However, the grouping behaviour of wild boar (Fernández–Llario et al., 1996) decreases the number of detections, and consequently, precise estimates of their population are particularly difficult to calculate. To yield a sufficient number of sightings and adequately estimate population density, we could also improve the design by increasing the number of replicates by site or by increasing the route survey length (with limitation from the road network extent and increase risk of double–counting of animals). Designing the surveys according to weather conditions could also help to increase the number of sightings. Using roads may also inevitably involve crossing habitats with variable visibility due to edges or topography. Also, the proposed methodology could be improved by adding a determination of the length of the surveyed transects where visibility is obstructed, as achieved by Heydon et al. (2000). Conclusion This study is the result of a pilot survey that highlights the limitations and advantages of large–scale multi–species monitoring using thermal imagery and distance sampling. It emphasized the need to develop a design in which covariates that can bias population estimates are taken into account. Detection probability associated with covariates may play an important role in any counting method and should be taken into account in subsequent density estimation analyses. Road–based distance sampling counts and thermal imagers seem to offer opportunities for detecting and monitoring large game species with patterns of nocturnal activities. However, for more elusive species such as wild boar, a modified monitoring survey should be designed in order to collect sufficient sightings to fit a detection function. As the random placement of transects is not possible, the proposed stratified–random selection method (random selection of suitable transects according to criteria) can help make roads a possibility for large–scale surveys while limiting potentially associated sources of bias. To our knowledge, this study is also the first to compare two types of thermal imager that differ in terms of resolution, portability and cost, and it showed that detectability was similar for both devices, even across highly contrasted habitat conditions. The cost of thermal imaging apparatus has likely prevented wider use of such devices in the field, but now that they are becoming more affordable their use in wildlife ground counts can be expected considering the numerous advantages they offer. However, although such advances may assure robust, valid estimates of animal population size, they will not obviate the need for a well–prepared survey design.

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Acknowledgements This research was supported by the Fonds pour la Formation à la Recherche dans l’Industrie et l’Agronomie (FNRS–FRIA, Belgium). We would like to thank the companies Anatherm SPRL and FLIR Ltd. Particularly Karinna Robe and Jan Delye for the thermography training and the use of the FLIR HS–324 thermal imager. Thanks also to the whole working team, Cedric Geerts, Marie Lebourges, Coralie Mengal, Alain Monseur, Adrien Schot, for their night availability and the good working atmosphere during the field work. References Acevedo, P., Ferreres, J., Jaroso, R., Duran, M., Escudero, M. A., Marco, J. & Gortazar, C., 2010. Estimating roe deer abundance from pellet group counts in Spain: An assessment of methods suitable for Mediterranean woodlands. Ecological Indicators, 10: 1226–1230. Acevedo, P., Ruiz–Fons, F., Vicente, J., Reyes– Garcıa, A. R., Alzaga, V. & Gortazar, C., 2008. Estimating red deer abundance in a wide range of management situations in Mediterranean habitats. Journal of Zoology, 276: 37–47. Anderson, D. R., 2001. The Need to Get the Basics Right in Wildlife Field Studies. Wildlife Society Bulletin, 29: 1294–1297. Apollonio, M., Andersen, R. & Putman, R. (Eds.), 2010. European ungulates and their management in the 21st century, Cambridge. Barancekova, M., Krojerova–Prokesova, J., Sustr, P. & Heurich, M., 2010. Annual changes in roe deer (Capreolus capreolus L.) diet in the Bohemian Forest, Czech Republic/Germany. European Journal of Wildlife Research, 56: 327–333. Barbraud, C. & Thiebot, J.–B., 2009. On the importance of estimating detection probabilities from at–sea surveys of flying seabirds. Journal of Avian Biology, 40: 584–590. Barrio, I. C., Acevedo, P. & Tortosa, F. S., 2010. Assessment of methods for estimating wild rabbit population abundance in agricultural landscapes. European Journal of Wildlife Research, 56: 335–340. Baubet, E., Bonenfant, C. & Brandt, S., 2004. Diet of the wild boar in the french alps. Galemys, 16: 101–113. Baubet, E., Brandt, S. & Fournier–Chambrillon, C., 2009. La consommation de vers de terre par le sanglier: quelle relation avec les dégâts sur prairie? Faune Sauvage. Baubet, E., Ropert–Coudert, Y. & Brandt, S., 2003. Seasonal and annual variations in earthworm consumption by wild boar (Sus scrofa scrofa L.). Wildlife Research, 30: 179–186. Boitani, L., Mattei, L., Nonis, D. & Corsi, F., 1994. Spatial and activity patterns of wild boars in Tuscany, Italy. Journal of Mammalogy, 75: 600–612. Bozec, Y.–M., Kulbicki, M., Laloe, F., Mou–Tham, G. & Gascuel, D., 2011. Factors affecting the detection

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Non–invasive genetic approaches for estimation of ungulate population size: a study on roe deer (Capreolus capreolus) based on faeces C. Ebert, J. Sandrini, B. Spielberger, B. Thiele & U. Hohmann

Ebert, C., Sandrini, J., Spielberger, B., Thiele, B. & Hohmann, U., 2012. Non–invasive genetic approaches for estimation of ungulate population size: a study on roe deer (Capreolus capreolus) based on faeces. Animal Biodiversity and Conservation, 35.2: 267–275. Abstract Non–invasive genetic approaches for estimation of ungulate population size: a study on roe deer (Capreolus capreolus) based on faeces.— Estimating population size is particularly difficult for animal species living in concealing habitats with dense vegetation. This is the case for roe deer as for many other ungulates. Our objective was to develop a non–invasive genetic capture–mark–recapture approach based on roe deer faeces collected along transects. In a pilot study, we collected 1,790 roe deer faeces during five sampling days in a forested study area in south western Germany. We extracted DNA from 410 of these samples and carried out microsatellite analysis using seven dinucleotide markers. The analyses resulted in 328 useable consensus genotypes which were assigned to 174 individuals. The population size estimated using a Bayesian approach was 94 (82–111) male and 136 (121–156) female roe deer. Our study shows that non–invasive genetic methods are a valuable management tool for roe deer. Key words: Abundance, Ungulates, Faecal pellets, Capreolus capreolus, Capture–mark–recapture, DNA. Resumen Estudios genéticos no invasivos, para la estimación del tamaño de una población de ungulados: estudio sobre el corzo (Capreolus capreolus) basado en sus heces.— La estimación de los tamaños de población es particularmente difícil en las especies de animales que viven en hábitats de vegetación densa, en la que se pueden mimetizar. Este es el caso del corzo, al igual que el de muchos otros ungulados. Nuestro objetivo fue desarrollar una aproximación genética no invasiva de captura–marcado–recaptura basada en las heces de corzo recogidas a lo largo de transectos. En un estudio piloto, recogimos 1.790 heces de corzo durante cinco días de muestreo en un área de estudio boscosa en el sudoeste de Alemania. Extrajimos el ADN de 410 de dichas muestras y llevamos a cabo un análisis de microsatélites utilizando siete marcadores de dinucleótidos. Los análisis tuvieron como resultado 328 genotipos consenso, que se asignaron a 174 individuos. La población estimada usando el enfoque bayesiano fue de 94 (82–111) machos y 136 (121–156) hembras. Nuestro estudio demuestra que los métodos genéticos no invasivos constituyen una herramienta de gestión muy valiosa para el corzo. Palabras clave: Abundancia, Ungulados, Heces, Capreolus capreolus, Captura–marcado–recaptura, ADN. Received: 11 I 12; Conditional acceptance: 10 II 12; Final acceptance: 3 IV 12 Cornelia Ebert, Bettina Spielberger & Bernhard Thiele, Seq–IT GmbH und Co KG, Pfaffplatz 10, 67755 Kaiserslautern, Germany.– Julian Sandrini & Ulf Hohmann, Research Inst. for Forest Ecology and Forestry Rhineland– Palatinate, Hauptstrasse 16, 67705 Trippstadt, Germany. Corresponding author: Cornelia Ebert. E–mail: cebert@gmx.de

ISSN: 1578–665X

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Introduction Assessing population size is of prime importance for the management of large herbivores such as the different deer species (Gordon et al., 2004; Williams et al., 2002). However, for elusive animal species living in habitats with dense vegetation, estimating population size is particularly difficult. This is the case for roe deer (Capreolus capreolus), and other ungulates living in concealing habitats (Smart et al., 2004; Hewison et al., 2007). For these species, indirect methods such as pellet counts seem more appropriate than direct counts, even though the latter have been applied rather successfully in deer populations in open areas (Smart et al., 2004; Tsaparis et al., 2009). Nevertheless, most traditional indirect approaches to estimate abundance yield estimates of low accuracy and low precision (Garel et al., 2010), and thus have limited power for certain management aspects. Direct methods such as traditional capture–mark–recapture have also been applied for roe deer, but mostly in enclosed study areas or in combination with radiotelemetry (Gaillard et al., 1986; Pegel & Thor, 2000; Hewison et al., 2007). In this context, non–invasive genetic methods based on hair or faeces represent a promising alternative to traditional methods because they can combine the advantages of indirect methods with the accuracy and precision of CMR approaches (McKelvey & Schwartz, 2004; Beja–Pereira et al., 2009). In the early 1990s, this approach was mainly used for carnivores such as coyotes and bears (e.g. Kohn et al., 1999; Woods et al., 1999), but it was not long before it was also used for studies on other species (reviewed in Waits & Paetkau, 2005; Beja–Pereira et al., 2008; Ebert et al., 2010). However, it is only in recent years that non–invasive genetic population size estimation methods —mostly based on faeces as a DNA source— have been carried out for ungulate species, such as Argali sheep (Ovis ammon, Harris et al., 2010), Sitka black–tailed deer (Odocoileus hemionus sitkensis; Brinkman et al., 2011), and wild boar (Sus scrofa; Ebert et al., 2009, in press). The roe deer is a widespread species in many parts of Europe and when it occurs in high densities it can have a significant impact on the vegetation in its habitat through browsing (Ellenberg, 1978; Gordon et al., 2004). In the Palatinate Forest, where this study was carried out, roe deer are managed mainly by hunting. However, in our study area, as in most regions, reliable data on roe deer abundance that would facilitate roe deer management (e.g. for setting appropriate harvest quotas) are lacking. We therefore aimed to establish a non–invasive genetic population estimation method for roe deer and to test its feasibility for this species. In this article, we present a pilot study based on faeces samples collected along transects. Material and methods Study area and faeces sampling Roe deer faeces sampling was carried out in a 2,400 ha segment of a wildlife research area situated in the Pa-

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latinate Forest in Rhineland Palatinate, south western Germany (49° 12′ N, 7° 45′ E). This wildlife research area (approx. 100 km²) was established in 2005. It is situated in the German part of the transfrontier biosphere reserve 'Vosges du Nord–Pfälzerwald', located along the French – German border. Forest cover in the study area is 93%, with nutrient–poor sandstone as the prevalent soil. The dominant tree species are Fagus sylvatica (33%), Pinus sylvestris (30%), Quercus spec. (16%), Picea abies (10%) and Pseudotsuga menziesii (8%) (Forestry of Rhineland–Palatinate, pers. comm.; Hohmann & Huckschlag, 2010). Several small settlements with surrounding open areas lie at the periphery of the study area. Altitude ranges between 220 and 611 m a.s.l. Annual average temperature is 8–9°C (Weiß, 1993), and annual precipitation approximates 600–1,000 mm. Besides roe deer, red deer (Cervus elaphus) and wild boar (Sus scrofa) are common ungulate species in the study area. The state forestry authorities in Rhineland–Palatinate manage the state–owned forests and are responsible for the hunting in these areas. The roe deer hunting bag in the forestry of Hinterweidenthal, in which our study area is located, is on average 2.0 animals/km2 (average between 1999 and 2009). The hunting season for roe deer is from May until January, with different age classes being allowed to be hunted at different times of the season. Hunting is performed by drive hunts (≥ 1,000 ha) between October and January and by single hunt throughout the rest of the season. We implemented 20 transects in a study area of approximately 4,000 ha (fig. 1). We aimed to cover the area as systematically as possible and to include every habitat type in the sampling scheme. Transect lengths varied between 4.9 and 7.2 km with an average length of 5.7 km. Distances between transects ranged between 231 m and 500 m with an average of 322 m. Each transect was searched by one person on each of five days (14–18 III 2011). The field workers followed transect routes using maps and compasses. The locations of all detected roe deer faeces were recorded using GPS loggers (Mobile Action Inc., I–gotU GT 120, http://www.i-gotu.com). Approximately one hand full of pellets of each detected pellet group was collected using an inverted freezer bag which was then reversed and closed. Samples were stored frozen (–19ºC) in the sealed freezer bags until analysis. The remaining pellets from each sampled pellet group were removed to avoid double sampling of the same faeces. Because of limited lab resources we were not able to analyze all collected faecal samples and thus had to select a subsample among the collected faeces. We selected samples proportional to the number of detected faeces in each transect. We selected approximately 20% of all samples in each of the 20 transects. The number of detected samples was unequally distributed over the transect grid. To guarantee a maximum chance of representative animal detection along the whole transect, we divided each transect in six subtransects of equal length. Next we selected at least one sample in each subtransect if possible. The residual samples were selected in proportion to the number of findings within the subtransects. When choosing among sam-

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United Kingdom Netherlands

Belgium

Germany

Luxembourg Study area

France

Genotyped samples Transcts

Wildlife Research Area

2.5

5 km

Fig. 1 Location of the study area in south western Germany and transect design. The size of the area covered by the transects within the 100 km² wildlife research area is approximately 40 km². Fig. 1. Localización del área de estudio en el sudoeste de Alemania y distribución de los transectos. El tamaño del área cubierta por los transectos, situada dentro de una zona de investigación de 100 km² en la naturaleza, es de aproximadamente 40 km².

ples on a subtransect, we used freshness as selection criterion in order to optimize the genotyping success. DNA extraction and genotyping DNA was extracted from the faeces samples within four weeks after collection using the NucleoSpin soil kit (Macherey–Nagel, Düren, Germany) following the manufacturer’s instructions, but with one modification: for the first step, three whole faecal pellets were vortexed together with one ml of lysis buffer for 15'. Seven dinucleotide microsatellites were analyzed to identify individual roe deer (table 1). To determine the sex of the sampled animals, we additionally amplified the Amelogenin gene according to Gurgul et al. (2010). A sample was classified as male when the Y–allele was present at least once. All samples showing only the X–allele in all three repeats were classified as female. For microsatellite analysis, all markers were combined in one multiplex reaction and a GeneAmp® PCR System 9700 Cycler (Applied Biosystems, Darmstadt, Germany) was used at the following PCR conditions: initial denaturation at 95°C for 15 min followed by 45 cycles of 30 s at 94°C, 30 s of 57°C and 60 s at 72°C, and a terminal elongation step at 60°C for 30 min. Amplification reactions were performed in triplicate, each in a total volume of 12 μl using the Qiagen Multiplex PCR kit (Qiagen, Hilden, Germany). We used the primers at concentrations of 0.2 μM to 0.4 μM. Amplification products were run on

an ABI3730 Sequencer using the ABI GS500LIZ size ladder and analyzed using the software GeneMapper v3.7 to determine allele lengths (Applied Biosystems, Darmstadt, Germany). We deduced consensus genotypes from the triplicate results. Samples were typed as heterozygous at one locus if both alleles appeared at least twice, and as homozygous when all replicates showed the same result. We repeated the genotyping another three times in cases when results were ambiguous after the first three replicates. All samples which failed to amplify or to produce unambiguous results (i.e. both alleles present at least two times for heterozygotes and only one allele in all six repeats for homozygotes) for more than two loci were discarded. We scrutinized genotypes differing by one (1–MM) or two (2–MM) alleles to detect genotyping errors. For all 1– or 2–MM pairs, raw data were re–checked to resolve the mismatches. Genotype pairs with only one mismatch were regarded as originating from the same individual (Ruell et al., 2009), whereas 2–MM pairs were considered as originating from different individuals, if re–checking of raw data and an additional three PCR repeats did not alter the results and if both samples matched with other samples in the data set (Paetkau, 2003). Determination of matching genotypes and construction of capture histories were carried out using GENECAP (Wilberg & Dreher, 2004). To confirm the power of the used loci, we calculated the probability of identity (PID) and, being more conservative, PID

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Table 1. Microsatellite markers used for individual identification of roe deer: conc. Concentration; N. Number of alleles. Tabla 1. Marcadores de microsatélites utilizados para la identificación individual de corzos: conc. Concentración; N. Número de alelos. Marker

Primer dye

Primer conc. (μM)

Length (bp)

Reference

Roe8

HEX

0.4

59–80

Vial et al., 2003

Roe6

6–FAM

0.2

87–105

Vial et al., 2003

MAF70Q

6–FAM

0.2

123–166

Vial et al., 2003

BMC1009

NED

0.2

276–291

Galan et al., 2003

BM848

VIC

0.4

353–367

Galan et al., 2003

BM757

6–FAM

0.2

175–210

Galan et al., 2003

VIC

0.4

158–193

Galan et al., 2003

OarFCB304

for siblings (PIDsibs; Waits et al., 2001) using GIMLET (Valière, 2002). Expected and observed heterozygosity as well as Hardy–Weinberg tests were performed using CERVUS 3.0 (Kalinowski et al., 2007). We calculated genotyping error rates (allelic dropout [ADO] and false alleles [FA]) as recommended in Broquet & Petit (2004). To illustrate the discriminating power of the loci used, we furthermore computed a match probability by multiplying the PID over all loci by an assumed approximate population size of 300 roe deer. As a blind test for genotyping reliability, we reanalysed 18 faeces samples without the lab staff knowing to which of the already analysed samples these further 18 samples should match. Furthermore, DNA was extracted and genotyping was carried out from tissue samples of all roe deer hunted in the study area after our sampling. The resulting genotypes were compared to those obtained via faeces sampling. Population estimation We calculated population size estimates using maximum likelihood mixture models for closed captures with heterogeneity (Pledger, 2000) implemented in Program MARK (White & Burnham, 1999). These models are widely applied in non–invasive genetic studies and we used them to facilitate comparison. Each of the five sampling days was considered as a separate ‘capture’ session. We defined a set of plausible candidate models with varying assumptions concerning capture probability (p) and recapture probability (c): (1) model p(.) = c(.) with constant capture– and recapture probability; (2) model p(.), c(.) accounting for behavioural response to sampling; (3) model pt = ct with capture– and recapture probabilities varying over time; and (4) model p1 = c1, p2 = c2 with two different mixtures for capture and recapture probabilities accounting for individual heterogeneity For each of the four basic models, we considered two different cases: 'basic model only' and 'basic

model including sex', in which the sex of the animals is included as a grouping factor (table 3). For population estimation, only those samples of each individual which were detected on different days were treated as 'recaptures', so that only one recapture per day was taken into account in order to fit into the traditional CMR framework (Otis et al., 1978). The different models were compared and ranked using Akaike’s Information Criterion with an additional bias correction term (AICc; Burnham & Anderson, 2002). Additionally, we calculated model averages, i.e. weighted averages over all models according to their support in the data as indexed by the AICc weights, in order to account for model selection uncertainty (Burnham & Anderson, 2002). Furthermore, we calculated CI using the unconditional standard error (SE) and the equation reported in Rexstad & Burnham (1992, page 19), because confidence intervals (CI) in Program MARK for model averages do not account for the minimum number of observed individuals. We additionally aimed to obtain CI for the total population (male + female). When sex is included as a grouping variable, as in our analysis, population sizes and CI are estimated separately for both sexes. We therefore calculated the sum of a random number of the female and male probability distribution, iterated this 10,000 times and calculated mean (total population size) and standard error from the resulting distribution. We used mean and standard error to calculate 95% CI for the estimated total population size based on the corresponding Rexstadt & Burnham (1992) equations. In addition to the closed captures estimates based on mixture models, we calculated population estimates using a Bayesian model (Gazey & Staley, 1986; Petit & Valière, 2006) This single session approach is especially suitable for non–invasive data because it can make full use of all detections for each individual in the data set. We examined the assumption of capture homogeneity by carrying out a test in

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Table 2. Characteristics of the microsatellite markers used for individual identification of roe deer (Hexp. Expected heterozygosity; Hobs. Observed heterozygosity; p–value. Bonferroni–corrected p–values for deviations from Hardy–Weinberg expected genotype frequencies; PCR+. Percent positive PCR. (Mean error rates [ADO and FA] were calculated after Broquet & Petit, 2004). Tabla 2. Características de los marcadores de microsatélites utilizados para la identificación individual de los corzos (Hexp. Heterocigosidad esperada; Hobs. Heterocigosidad observada; p–values. Valores de probabilidad con la corrección de Bonferroni para las desviaciones de las frecuencias genotípicas esperadas de Hardy–Weinberg; PCR+. PCR con porcentaje positivo. (Las tasas de error medio [ADO y FA] se calcularon según Broquet & Petit, 2004).

Marker Roe8

Hexp 0.71

Hobs 0.71

p–value

PCR+

ADO

0.81

0.98

0.028

0.000

Roe6

0.68

0.67

0.95

0.93

0.046

0.000

MAF70Q

0.77

0.76

0.45

0.99

0.048

0.000

BMC1009

0.66

0.65

0.80

0.97

0.077

0.000

BM848

0.71

0.67

0.17

0.94

0.064

0.000

BM757

0.86

0.88

0.10

0.97

0.043

0.000

OarFCB304

0.84

0.77

0.08

0.97

0.033

0.000

Mean

0.74

0.73

0.96

0.048

0.000

which the sampling process is simulated under the assumption of homogeneity and the expected number of captures is compared with the observed number of captures per individual (Puechmaille & Petit, 2007). The Bayesian estimate and the test were performed using the R package (Ihaka & Gentleman, 1996) and a script provided by E. Petit (pers. comm.). Results Faeces sampling and genotyping During the five sampling days, 1790 roe deer faeces were collected of which we selected 410 samples in the subsampling process for DNA extraction and microsatellite analysis. Of these, 328 samples (i.e. 80%) yielded a useable consensus genotype consisting of at least 5 markers. For fifteen samples, one marker failed to amplify and for four samples two markers were missing. The proportion of positive PCR varied among loci between 0.93 and 0.98 (table 2). Mean expected heterozygosity (Hexp) was 0.75 and mean observed heterozygosity (Hobs) was 0.73; no significant deviations from Hardy–Weinberg–equilibrium were detected (table 2). The observed PID was 2.16 x 10–8, and PIDsibs was estimated as 0.0012. The ADO rate varied between markers with a minimum of 0.028 (Roe8), a maximum of 0.077 (BMC1009) and an average of 0.048 (table 2). No false alleles were detected in the data set. The match probability was calculated as 6.48 x 10–6. After re–checking the genotypes, no 1–MM pairs and seven 2–MM pairs

remained in the data set. All samples of the 2–MM pairs matched with at least one other sample and were thus considered to belong to different individuals. The 328 genotypes were assigned to 174 different individuals, 71 males and 103 females. Of the 174 individuals, 89 were sampled once, 46 were sampled twice, 16 were sampled three times, 16 were sampled four times, 6 were sampled five times, and one was sampled six times. All 18 samples reanalyzed in the blind test matched the genotype of the original sample. Tissue samples from 31 hunted roe deer were genotyped and resulted in 18 matches with genotypes already known from the faeces sampling. Population estimation Four of the maximum likelihood models received considerable support with ∆AICc < 2 (table 3). The most supported model incorporated time and sex dependent capture probabilities, whereas the time dependent model, the heterogeneity model and the model with constant p and c were less supported. Mean p was 0.234 for male and 0.256 for female roe deer. Estimated p over the five sampling days were 0.25, 0.22, 0.17, 0.28 and 0.25 for males and 0.32, 0.29, 0.28, 0.25 and 0.14 for females (fig. 2). The estimated population size ranged from 89 to 113 male roe deer and from 118 to 164 female roe deer with confidence interval widths varying between 25% and 90% of the estimated population sizes for the four most supported models (table 3). The model averaged population size was 99 (80–155) male and 139 (114–218) female roe

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Table 3. Roe deer population estimates derived from non–invasive genetic sampling data based on faeces. The maximum likelihood mixture models are ranked according to Akaike’s Information Criterion (AICc). Capture probability is symbolized by p, recapture probability by c. A dot after a parameter denotes that the parameter is held constant in the respective model, a 't' stands for a parameter varying over time. Further parameters are K (number of parameters of each model), wi (model weight), N (estimated population size) for each sex incl. 95% confidence intervals and standard error (SE). 'Sex' models include the sex of the animals as a grouping factor. Tabla 3. Estimaciones de la población de corzo derivadas del muestreo de datos genéticos no invasivo, basado en las heces. Los modelos mixtos de máxima probabilidad se ordenaron según el Criterio de Información de Akaike (AICc). La probabilidad de captura se simboliza mediante una p, la probabilidad de recaptura, mediante una c. Un punto tras un parámetro significa que dicho parámetro se mantiene constante en el modelo respectivo, una ''t'' se refiere a un parámetro que varía con el tiempo. Otros parámetros son K (el número de parámetros de cada modelo), wi (peso del modelo), N (tamaño de la población estimado) para cada sexo incl. 95% de intervalos de confianza y error estándar (SE). Los modelos ''sex'' incluyen el sexo de los animales como factor de agrupamiento. Model

AICc

ΔAICc

Male K

N (95% CI)

Female N (95% CI)

pt = ct, sex

–186.97

0.000 0.348

96 (84–119)

8.72

133 (119–156) 8.90

pt = ct

–185.65

1.321 0.179

93 (84–110)

6.28

136 (123–157) 8.53

p1 = c1, p2 = c2

–185.56

1.409 0.172

113 (86–190) 24.06 164 (125–273) 34.26

p(.) = c(.)

–185.28

1.694 0.149

94 (84–111)

6.70

136 (123–158) 8.62

p(.), c(.)

–183.70

3.273 0.068

89 (79–113)

8.26

129 (115–162) 11.33

p(.), c(.), sex

–183.32

3.654 0.056

115 (83–232) 32.66 118 (109–143) 8.03

p1 = c1, p2 = c2, sex

–180.83

6.138 0.016

116 (89–183) 22.27 168 (113–533) 80.52

p(.) = c(.), sex

–179.47

7.492 0.008

96 (84–120)

8.83

–

99 (80–155)

16.97 139 (114–218) 23.33

Model average

–

deer, leading to an estimated total population size of 238 (201–323) roe deer (SE = 28.83). When taking into account the most supported model, the total population size was 229 (209–259) roe deer (SE = 12.45). The Bayesian population estimate was 94 (82–111) for male roe deer and 136 (121–156) for female roe deer. No heterogeneity in capture probability was found for the female part of the population, whereas for the males, the heterogeneity test revealed that the observed capture frequencies differed significantly from the expected capture frequencies, thus indicating heterogeneity. Discussion Considering that 80% of the samples yielded a useable consensus genotype after three to six PCR repeats, the DNA extraction and genotyping protocol can be regarded as efficient and the PCR success rate as rather high (Broquet et al., 2007). Although several genotyping errors occurred, mainly in the form of ADO, we believe the overall misidentification rate to be low because ambiguous samples were

134 (120–158) 9.19

repeated up to six times and error rate was below 5% (Lukacs & Burnham, 2005). Furthermore, all samples which were analysed twice in the blind test resulted in matching genotypes, and 18 of 31 roe deer which were harvested in the study area after our sampling trial had genotypes that matched individuals which were detected via faeces sampling. This indicates that the genotyping protocol is repeatable and consistent. As the PID, the match probability and, particularly, the PIDsibs (being well below 0.01) were low, the set of markers seems sufficient to reliably discriminate even closely related individuals from each other (Woods et al., 1999). Mean capture probabilities (p) were consistently above 0.2 with a somewhat higher p for the female part of the population than for males (0.256 and 0.234, respectively). The sex ratios of the sample and the estimated population of between 1.4 and 1.45 can be considered realistic, since in many ungulate species populations tend to be female–biased (Clutton–Brock & McLonergan, 1994) and in a study in southern Germany based on direct counts the mean roe deer sex ratio was estimated as 1.5 (Pegel & Thor, 2000). The fact that the most supported maximum

likelihood model includes differences between the sexes in p suggests that males were slightly less prone to detection than females. This observation could perhaps be induced by differences in space or habitat use between the sexes (Mysterud, 1999). Both sexes showed a certain variation of p over time. As p did not decrease significantly over time for males a reaction of the animals to a disturbance by the field workers is not probable. For the females, however, p declined slightly over the five sampling days with a more distinct drop at day five (fig. 2). In the first sampling days the sample size is often higher because faeces have accumulated in the days before the beginning of the trial, and a decline after the first sampling days is therefore to be expected. However, several studies suggest that female ungulates are more wary towards human disturbance than males, which would explain the sex–related differences in p (Stankowich, 2008). Roe deer, nevertheless, tend to have rather short flight distances when disturbed by humans walking in their habitat (in mean between 39 and 85 m; De Boer et al., 2004), and thus it is unlikely that animals left the sampled area in reaction to the field workers. Brinkman et al. (2011) observed similar variations in capture probability for Sitka black–tailed deer and suggested variation in activity and habitat use during winter as a reason. The heterogeneity model (p1 = c1, p2 = c2) also received considerable support, suggesting the incidence of a certain individual heterogeneity (which in fact is almost ubiquitous in non–invasive as well as in traditional CMR studies; Rudnick et al., 2008). Garel et al. (2010) showed that changing observation conditions (e.g. weather) can also have a significant impact on observation results. In our case, this is unlikely as an explanation for the decrease in p over time or for individual heterogeneity, because weather and overall sampling conditions were good and did not change markedly over the five days of sampling. The Bayesian population estimates for both sexes are very similar to the maximum likelihood estimates but show higher precision. The estimated population sizes of the Bayesian approach and the supported maximum likelihood models are relatively consistent and show moderate precision (table 3), which together with reasonably high capture probabilities suggests that a representative coverage of the population has been achieved. However, in order to allow a quantitative evaluation of management measures, population densities will be needed in addition to population sizes. Therefore, the topics of population closure and density calculation should be addressed in future studies. The assumption of demographic closure can be met by choosing a short time span for sampling like we did in our study. In contrast, geographic closure can be very difficult to achieve and closure violation can severely bias density estimates (Obbard et al., 2010). This should be taken into account when calculating population densities. An alternative to traditional methods for population density calculation (e.g. using buffers to determine an effectively sampled area by mean maximum recapture distance, Parmenter et al., 2003) are spatially explicit capture–recapture appro-

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Capture probability

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0.4 0.3 0.2 0.1 0

2 3 4 Sampling day

Fig. 2. Estimated capture probabilities (p) for male and female roe deer during a non–invasive genetic faeces sampling study. Solid rhombi represent male roe deer and open circles represent females; error bars represent the standard error. Fig. 2. Probabilidades de captura estimadas (p) para corzos machos y hembras durante un estudio de muestreo genético no invasivo basado en las heces. Los rombos negros representan a los machos y los círculos blancos a las hembras; las barras de error representan el error estándar.

aches (SECR), which hold the potential to mitigate the problem of closure violation in density estimates (Borchers & Efford, 2008). Furthermore, it would be an interesting topic for further research to evaluate the performance of the non–invasive genetic approach in comparison with traditional methods such as pellet counts or distance sampling. Even though the precision of our roe deer estimates seems to be good compared to many traditional estimates for woodland deer (as discussed e.g. in Cederlund et al., 1998; Smart et al., 2004), a detailed comparison between approaches remains very difficult. Different approaches need to be applied in the same study area and ideally in a population of known size in this context. Our study shows that non–invasive genetic sampling is a promising tool for roe deer management and suggests it could play a useful role in calibrating measures such as harvest quotas. Acknowledgements We would like to thank all those who helped collecting samples. Special thanks are due to Florian Feind for very valuable help in the field and Theresa Thiele for assistance with the sampling data. We also thank Felix Knauer for advice with the confidence interval calculations and Christophe Bonenfant and François Klein for fruitful discussion. Furthermore, we are indebted to the Forestry of Hinterweidenthal for helpful cooperation throughout the study.

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landscape using non–invasive capture–recapture sampling. Journal of Mammalogy, 90: 129–135. Smart, J. C. R., Ward, A. & White, P. C. L., 2004. Monitoring woodland deer populations in the UK: an imprecise science. Mammal Review, 34: 99–114. Stankowich, T., 2008. Ungulate flight responses to human disturbance: A review and meta–analysis. Biolgical Conservation, 141: 2159–2173. Tsaparis, D., Katsanevakis S., Ntolka E. & Legakis A., 2009. Estimating dung decay rates of roe deer (Capreolus capreolus) in different habitat types of a Mediterranean ecosystem: an information theoretic approach. European Journal of Wildlife Research, 55, 167–172. Valière, N., 2002. GIMLET: a computer program for analysing genetic individual identification data. Molecular Ecology Notes, 2: 377–379. Vial L., Maudet C. & Luikart G., 2003. Thirty–four polymorphic microsatellites for European roe deer. Molecular Ecology Notes, 3: 523–527. Waits, L. P., Luikart, G. & Taberlet, P., 2001. Estimating the probability of identity among genotypes in natural populations: cautions and guidelines. Molecular Ecology, 10: 249–256. Waits, L. P. & Paetkau, D., 2005. Noninvasive genetic sampling tools for wildlife biologists: a review of applications and recommendations for accurate data collection. Journal of Wildlife Management, 69: 1419–1433. Weiß, A., 1993. Pflege– und Entwicklungsplan Naturpark Pfälzerwald. Verein Naturpark Pfälzerwald e.V., Bad Dürkheim. [In German.] White, G. C. & Burnham, K. P., 1999. Program MARK: Survival estimation from populations of marked animals. Bird Study, 46 (supplement): 120–138. Wilberg, M. J. & Dreher, B. P., 2004. GENECAP: a program for analysis of multilocus genotype data for non–invasive sampling and capture–recapture population estimation. Molecular Ecology Notes, 4: 783–785. Williams, B. K., Nichols, J. D. & Conroy, M. J., 2002. Analysis and management of animal populations: modelling, estimation, and decision making. Academic Press, San Diego, California, USA. Woods, J. G., Paetkau, D., Lewis, D., McLellan, B. N., Proctor, M. & Strobeck, C., 1999. Genetic tagging of free–ranging black and brown bears. Wildlife Society Bulletin, 27: 616–627.

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Landscape ecology and wild rabbit (Oryctolagus cuniculus) habitat modeling in the Mediterranean region M. Narce, R. Meloni, T. Beroud, A. Pléney & J. C. Ricci

Narce, M., Meloni, R., Beroud, T., Pléney, A. & Ricci, J. C., 2012. Landscape ecology and wild rabbit (Oryctolagus cuniculus) habitat modeling in the Mediterranean region. Animal Biodiversity and Conservation, 35.2: 277–283. Abstract Landscape ecology and wild rabbit (Oryctolagus cuniculus) habitat modeling in the Mediterranean region.— Landscape modification is one of the reasons for the decrease in rabbit populations. The objective of this study was to model wild rabbit habitat using landscape ecology to create a diagnosis method able to assess habitat quality at a large scale. Rabbit presence/absence was recorded on 536 plots of 1 ha. Spotlight transect counts indicated a low relative abundance (KIA = 2.3 rabbits/km). We produced a land use map with metric precision using remote sensing. Water, bare soil, herbaceous, shrubs and trees were identified. Landscape structure and diversity were evaluated using variables available in FRAGSTATS. A logistic regression was performed to assess the link between rabbit presence/absence and landscape structure. Our results indicate that a suitable habitat has a high diversity, a medium number of patches and a small proportion of shrubs. These results could be used to diagnose the landscape prior to any management action to enhance rabbit populations and conversely be helpful as a tool of integrated control in the cases of local outbreaks with agricultural damages. Key words: Wild rabbit, Habitat modeling, Landscape pattern, Remote sensing, Mediterranean region. Resumen Ecología del paisaje y modelización del hábitat del conejo (Oryctolagus cuniculus) en la región mediterránea.— La modificación del paisaje es una de las razones de la disminución de las poblaciones de conejo. El objetivo de este estudio era la modelización del hábitat del conejo, usando la ecología del paisaje para crear un método de diagnóstico capaz de evaluar la calidad del hábitat a gran escala. Se observó la presencia/ausencia de conejos en 536 parcelas de 1 ha. Los recuentos nocturnos mediante transectos indicaron una abundancia relativa baja (IKA = 2,3 conejos/km). Por otra parte se realizó, mediante teledetección, una cartografía de uso del suelo con precisión métrica. Se identificaron el agua, el suelo desnudo, las plantas herbáceas, arbustos y los árboles. Se evaluaron la estructura del paisaje y su diversidad utilizando las variables disponibles en FRAGSTATS. Para evaluar la relación entre la presencia/ausencia del conejo y la estructura del paisaje se utilizó una regresión logística. Nuestros resultados indican que un hábitat adecuado tiene una gran diversidad, un número medio de parcelas y una pequeña proporción de arbustos. Estos resultados podrían ser utilizados para diagnosticar el paisaje antes de cualquier acción de gestión enfocada a aumentar las poblaciones de conejo, y a la inversa, ser un instrumento de control integrado en los casos de invasiones locales que causan daños a la agricultura. Palabras clave: Conejo, Modelización del hábitat, Estructura del paisaje, Teledetección, Región mediterránea. (Received: 22 XII 11; Conditional acceptance: 12 III 12; Final acceptance: 25 IV 12) Mathieu Narce, Timothée Beroud, Anne Pléney & Jean–Claude Ricci, Inst. Méditerranéen du Patrimoine Cynégétique et Faunistique, Domaine expérimental agri–environnement, Villa 'Les Bouillens', 30310 Vergèze, France.– Roberto Meloni, CEMAGREF, UMR TETIS, 34093 Montpellier, France. Corresponding author: Mathieu Narce. E–mail: instmed@impcf.fr ISSN: 1578–665X

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Introduction Populations of European wild rabbit (Oryctolagus cuniculus) have decreased in France since the 1950s (Marchandeau, 2000). This decline is mainly due to diseases (myxomatosis and rabbit viral haemorrhagic disease), evolution of agricultural practices and excessive hunting (Trout et al., 1992; Villafuerte et al., 1995; Marchandeau et al., 1998). Rabbits are considered a keystone species in the Mediterranean region (Delibes–Mateoset al., 2007). Because they are prey for many threatened species such as Bonelli’s eagle Aquila fasciata or eagle owl Bubo bubo (Delibes & Hiraldo, 1981) recovery of rabbit populations could have a positive effect on the conservation of these species (Moreno & Villafuerte, 1995). Furthermore, the rabbit was the principal game species in France until the 1950s and it continues to be a main target species To enhance rabbit populations, detailed knowledge of habitat requirements is fundamental. In recent years, many studies have been carried out along these lines (Rogers & Myers, 1979; Calvete et al., 2004; Carvalho & Gomes, 2004; Beja et al., 2007; Serrano Pérez et al., 2008) and several reports have pointed out the importance of the landscape structure. More specifically, they have observed that the interspersion of shelter and open areas is linked to the predation risk (Palomares & Delibes, 1997; Villafuerte & Moreno, 1997). The aim of this study was to create a diagnostic method able to assess habitat quality at a finer spatial scale than in previous works and in large areas (several hundred or thousand hectares). This could allow us to develop guidelines for managers and help them to increase rabbit populations. We first developed a high definition (metric precision) land–use map using remote sensing and automated photo–interpretation (Lucas et al., 2007). We then analysed the landscape structure to create a habitat suitability model in order to identify factors related to rabbit distribution.

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index of 2.3 rabbits/km was obtained (Ballinger & Morgan, 2002). It indicated a low relative abundance of rabbit. Determination of presence/absence areas Rabbit presence/absence was determined on 536 plots of 1 ha distributed in the various pilot areas. The plot area was close to the size of the home range of the rabbit (Lombardi et al., 2007; Marchandeau et al., 2007). Each plot was randomly prospected during 15’ noting each sign of rabbit presence: pellets, latrines, burrows and scratches. Two field sessions were conducted during the third quarter of 2008 and 2009 and two others during the first quarter of 2009 and 2010. Thus, each plot was prospected four times. The information was synthesized by considering that rabbits were present on a plot if their presence was noted at least once. Rabbits were considered absent if their absence was noted four times on a plot. Land use cartography In order to quantify landscape structure by calculating several metrics, a land use map with metric precision was performed using remote sensing technology. Aerial photographs with dot pitch of 50 cm were taken in July 2009 using the near infrared channel and the red, green and blue channel. An automated photo– interpretation was made in two steps with eCognition (Trimble) software (Baatz et al., 2001) using the object–oriented classification method which is suitable for very high resolution (Xiaoxia et al., 2005). First, we performed a segmentation consisting of delimiting polygons with homogenous land use. Each polygon was then classified as one of the land use types. The five different land use types identified were: water, bare soil, herbaceous (< 1 m), shrubs (1 << 7 m) and trees (> 7 m). Finally, the data were rasterized with cells of 4 m² (2 m x 2 m) (Xiaoxia et al., 2005).

Material and methods

Calculation of landscape metrics

Study area

Landscape structure and diversity were evaluated using 20 variables. Eleven were calculated using ArcMap™ 9.1 software using FRAGSTATS interface (McGarigal & Marks, 1994): Number of Patches (NUMP), Patch Richness (PR), Largest Patch Index (LPI), Mean Patch Size (MPS), Edge Density (ED), Area Weighted Mean Shape Index (AWMSI), Area Weighted Mean Patch Fractal Dimension (AWMPFD), Mean Nearest Neighbor (MNN), Interspersion Juxtaposition Index (IJI), Total Core Area (TCA) and Core Area Density (CAD). The width of the buffer needed to define the core area of the patches was set at 25 m. The nine others variables were directly calculated with ArcMap. PROPWATER, PROPBARESOIL, PROPHERB, PROPSHRUBS and PROPTREES indicate the proportion of the different land use types in the landscape. PROPOPEN and PROPCOVER, respectively, indicate the proportion of open areas (bare soil and herbaceous) and cover (shrubs and trees). NEARESTOPENAREA is the mean distance between points

The study area is located in the southeast of France and covers the Natura 2000 (ZPS) area of Durance valley (figs. 1A, 1B). It is 240 km long and 2 km wide (1 km on both sides of the river) for an area of 46,500 ha. The altitude varies from 15 m to 780 m a.s.l. and there is a Mediterranean climate (average annual temperature and rainfall, 12.4°C and 720 mm respectively) (Benichou & Le Breton, 1987). The land use in the study area is: 12% of artificial area, 38% of agricultural area, 42% of forest and semi–natural area and 8% of wetland and water (Union européenne– SOeS, 2006). The main natural areas are riparian forest with Populus sp., Alnus sp., Salix sp. and Quercus sp. Agricultural practices are mainly mixed farming with crop, arboriculture and breeding. Six pilot areas of about 100 ha each were demarcated and divided into plots of 1 ha. Using spotlight transect counts, an average kilometric abundance

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randomly positioned in cover areas and the nearest open area. For each plot of 1 ha, the surface of cover was calculated and one point was used for 100 m² of cover with a minimum of 30 points. The same was applied for the NEARESTCOVER which indicates the average mean distance to the nearest cover. All the variables are quantitative except PR which is qualitative with three modalities (one or two patch types: PR = a, three patch types: PR = b, four or five patch types: PR = c).

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Paris

Statistical analysis All the statistical analysis were performed using R© software (2.12.2 version) (R Development Core Team, 2011). The descriptive analysis reveals the presence of plots with very high values of NUMP and AWMSI. The specific study of these plots indicates that they were composed of orchards and row crops. Because of this particular landscape structure, remote sensing was poor. Thus, they were considered as outliers and removed from the data set which was finally composed of 516 measures. The correlation matrix indicates some colinearity and this was taken into account by computing the Variance Inflation Factor (VIF) with a threshold of 3 to test colinearity among the variables included in the different models. First, we compared the means of the variables from presence versus absence areas. Homoscedasticity was tested by an F–test. In cases of homoscedasticity, a t–test was performed. Otherwise, a Welch’s test was carried out. A generalized linear model was then used with Logit as a link function to assess the link between rabbit presence/absence (binary independent variable) and the landscape metrics (dependant variables) (Traissac et al., 1999). The data set was randomly divided in two parts: 80% of the sample for model calibration and the remaining 20% for validation. The first selection was done by minimizing Akaike Information Criterion (AIC) (Akaike, 1974; Lebreton et al., 1992; Posada & Buckley, 2004) with a stepwise selection starting with null model (function 'step' under R software with argument direction = 'both') (R Development Core Team, 2011). It consists of adding variables one by one and trying to remove those of the previous steps in order to minimize AIC. On the first model obtained by stepwise selection, Wald’s test was used to identify variables whose coefficient was not significantly different from zero (Hauck & Donner, 1977); these variables were removed. Finally, the sign of the coefficient of the different variables was examined. When we failed to find a logical and biological explanation (inconsistencies with other results in this study and with literature, no colinearity), the variable was removed. Interactions of first degree were tested with Wald´s test in the model obtained by stepwise selection and in the final model. The goodness of fit was tested with Hosmer–Lemeshow test (Lemeshow & Hosmer, 1982; Hosmer & Lemeshow, 2000) and McFadden’s pseudo r–squared. The discrimination capacity of the model was evaluated by means of a confusion matrix and Area Under the Receiver Operating Characteristic (ROC) Curve (Agresti, 2002).

Marseille 0

200

400 km

N 0

50 km

Fig. 1. Study area in the southeast of France (A) and location of the six pilot areas (B). Fig. 1. Área de estudio en el sureste de Francia (A) y localización de las seis áreas experimentales (B).

Results Mean difference between presence and absence areas are summarized in table 1. A significant difference was found for 15 variables. Seven variables had a higher value in the presence areas: IJI, NUMP, ED, AWMSI, AWMPFD, PR and PROPTREES. In contrast, MNN, MPS, TCA, CAD, LPI, PROPSHRUBS, NEARESTCOVER and NEARESTOPENAREA were significantly higher in areas of absence. The first model obtained by stepwise selection was composed of the following variables: NUMP, NUMP², PR, PROPSHRUBS, IJI and TCA. The VIF indicates that there was no colinearity among these variables. According to the Wald test, the coefficient of IJI was not significantly different from 0 (P = 0.151), so this variable was removed from the model. The positive coefficient of TCA seems to specify that large patches with an important core area would be more suitable for rabbit. This result was not due to colinearity and it was not consistent with other results of the present study or with the literature; so this variable was also removed. The final model was made up of four variables: NUMP,

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Table 1. Mean difference between variables in absence areas (Aa) vs. presence areas (Pa). (For abbreviations of variables see Material and methods. Tabla 1. Diferencias entre las medias de las variables en las zonas de ausencia (Aa) vs. presencia (Pa). (para las abreviaturas de las variables, ver Material and methods.)

Aa (N = 132)

Variables

Mean ± SD

P–value

MNN

6.84 ± 5.44

4.64 ± 2.17

1.28 x 10–05 ***

IJI

50.2 ± 30.0

65.9 ± 19.2

7.69 x 10–08 ***

NUMP

38.7 ± 37.2

65.7 ± 35.6

4.41 x 10–13 ***

MPS

0.06 ± 0.08

0.03 ± 0.03

2.43 x 10–07 ***

1,419.5 ± 758

1,939.8 ± 724

6.37 x 10–12 ***

AWMSI

2.31 ± 0.85

2.82 ± 0.84

3.97 x 10–09 ***

AWMPFD

1.19 ± 0.09

1.25 ± 0.07

1.13 x 10–09 ***

TCA

0.04 ± 0.07

0.01 ± 0.04

3.06 x 10–05 ***

CAD

61.4 ± 61.3

31.8 ± 59.9

1.50 x 10–06 ***

LPI

64.1 ± 23.8

53.7 ± 21.5

4.74 x 10–06 ***

3.22 ± 1.22

3.79 ± 0.84

1.20 x 10–07 ***

PROPSHRUBS

0.25 ± 0.26

0.17 ± 0.16

1.15 x 10–03 **

PROPTREES

0.20 ± 0.27

0.29 ± 0.26

4.65 x 10–04 ***

NEARESTCOVER

11.84 ± 11.3

8.64 ± 8.82

3.42 x 10–03 **

NEARESTOPENAREA

9.8 ± 11.49

7.3 ± 8.28

0.023 *

PROPHERB

0.42 ± 0.35

0.35 ± 0.25

0.063

PROPBARESOIL

0.11 ± 0.21

0.16 ± 0.23

0.068

PROPWATER

0.02 ± 0.05

0.02 ± 0.07

0.396

PROPOPEN

0.54 ± 0.34

0.52 ± 0.29

0.600

PROPCOVER

0.46 ± 0.34

0.48 ± 0.29

0.600

NUMP², PR and PROPSHRUBS (table 2). All of them were significant according to the Wald’s test. Interactions were tested in the various previous models but were not significant. Figure 2 (A and B) illustrates the effect of the variables on the probability of rabbit presence. Figure 2A indicates that the probability of presence is higher when there are three or more patch types in the landscape, suggesting it should be diversified to be favourable. Figure 2B summarises the effect of the proportion of shrubs and the number of patches on the probability of presence. The probability decreased when the proportion of shrubs increased. Because of the presence of the square term of NUMP, the probability of presence correlated positively with the number of patches up to about 100 patches/ha, after which it was negatively correlated. It thus reaches its maximum when there are about 100 patches per hectare.

Pa (N = 384)

Table 2. Coefficients of the variables retained in the model and P–value of Wald test. Tabla 2. Coeficientes de las variables utilizadas en el modelo y P–value del test de Wald. Term

Estimates

P–value Wald

Constant

–1.841

3.6 x 10–4 ***

NUMP

0.066

4.12 x 10–8 ***

NUMP²

–3.58 x 10–4 2.76 x 10–6 ***

PR = b

1.323

1.92 x 10–2 *

PR = c

1.339

1.54 x 10–2 *

PROPSHRUBS

–1.756

3.26 x 10–3 **

Animal Biodiversity and Conservation 35.2 (2012)

A Response of patch richness

281

Response surface for PR = b 0.8

0.85

ility Probabence of pres

0.6

0.80

0.4 0.2

0.75

0. 8

0. 4

b Patch richness

0. 6

0.70

0 10

0. 2

Pr of op sh or t ru ion bs

0 15

Probability of presence

0.90

r be s m tche u N pa of

Fig. 2. Relationship between probability of presence and patch richness (A) and relationship between probability of presence, proportion of shrubs and number of patches for PR = b (B). Fig. 2. Relación entre la probabilidad de presencia y la diversidad de parcelas (A) y relación entre la probabilidad de presencia, la proporción de arbustos y el número de parcelas para PR = b (B).

The Hosmer–Lemeshow test (3.99, P = 0.86) indicated that the model correctly fitted the data. McFadden’s pseudo r–square (0.19) was also computed and indicated an adequate fit. The Variance Inflation Factor revealed no colinearity. Using the validation data set, the area under the ROC curve (AUC) was 0.79. This value shows that the model had a high capacity for discrimination. Sensitivity and the specificity were calculated using the confusion matrix and were 76% and 73%, respectively. Discussion The aim of this study was to identify landscape factors that influence rabbit distribution. Climatic conditions and altitude were thus not used as predictor variables. However, even if there is a large range of altitude on the study area, climatic conditions are consistent with rabbit requirements. To detect the presence of rabbits, all the signs of presence were taken into account rather than only the traditionally used pellets. The risk of declaring an absence where the rabbit was present was therefore minimal. The endpoint was to identify suitable habitats on the basis of rabbit distribution. Their distribution also depends on other factors, however, such as disease impact, and it can vary from one year to the next. To minimize these biases, four field sessions were conducted, spread over three years and an area was considered as suitable if rabbit presence was recorded at least once. However, the

absence of rabbit from some suitable areas cannot be totally ruled out. The probability threshold used to discriminate rabbit presence and absence was set at 0.75. This value permitted to maximise the discrimination capacity of the model. It only discriminated between suitable and unsuitable habitats because it works as a binary variable. However, when the model is used for prediction, it can be modified according to the risk to be minimised. Increasing the threshold would limit the number of false positive cases and decreasing it would limit false negative cases. In this study, the land use map was created using remote sensing which combines high definition aerial photography with automated photo–interpretation. Aerial photography with a dot pitch of 50 cm allowed us to make a more precise map than those using classical aerial photography or 1:25,000 maps. Thus, the map was composed of pixels of 4 m² which seem to be consistent with a rabbit habitat study. This high precision could be interesting to precisely evaluate habitat structure in relation to the perception scale of the rabbit. Moreover, the use of remote sensing can work on very large areas. This technique avoids manual digitalisation which is very time consuming on areas of several hundred hectares. However, it has a disadvantage because it only takes into account the highest vegetation layer. For example, it ignores the presence of herbaceous or shrubs under trees. The results presented in table 1 suggest that fifteen variables could be useful to describe rabbit habitat,

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even if only four of them are used in the final model. The variables IJI, AWMSI and AWMPFD underline the importance of the fractal dimension (Jimenez et al., 2006) and interspersion (Carvalho & Gomes, 2004) between patches. The statistical significance of the variables NUMP, PROPSHRUBS and NEARESTCOVER indicates that a suitable habitat is a patchwork with some evenly distributed shrub cover. This is consistent with the results of Moreno & Villafuerte (1995) that indicated that rabbit abundance was greatest at less than 20 m from scrub cover. The importance of the arrangement of cover and open areas can be considered an anti– predator strategy. Cover is safe for the rabbit to hide from the birds of prey during the day. On the contrary, open areas are safer at night because carnivorous mammals use cover for hunting (Moreno et al., 1996). This study provides some characteristics of rabbit habitat obtained using remote sensing using a fine spatial scale. Based on our results, we have a diagnostic method which could be used to assess habitat quality concerning rabbit requirements. This process could enable accurate discrimination between suitable and unsuitable areas over areas of several hundred or thousand hectares. Evaluation of unsuitable areas might identify variables of inadequate value. Such information would be of valuable help for managers trying to improve habitat quality in order to enhance rabbit populations. Conversely, this method could be used as a tool of integrated control to reduce rabbit populations (Boag, 1987). In the case of local outbreaks causing agricultural damage, modification of habitat structure could create an unsuitable habitat and lead to a long–term decrease of rabbit populations. Acknowledgements We would like to thank all colleagues and those who contributed to this study, especially the Fédérations Départementales de Chasseurs (04, 05, 13, 83, 84) for technical services and the Fédération Nationale des Chasseurs (FNC) and the Fédération Régionale des Chasseurs de Provence–Alpes–Côte–d’Azur (FRC PACA) for financial support. We are also indebted to an anonymous referee for helpful comments and critical review of the manuscript. References Agresti, A., 2002. Categorical data analysis. 2nd edn. John Wiley & Sons, Inc., Hoboken. Akaike, H., 1974. A new look at the statistical model identification. IEEE Transactions on Automatic Control, 19: 716–723. Baatz, M., Benz, U., Dehghani, S., Heynen, M., Höltje, A., Hofmann, P., Lingenfelder, I., Mimler, M., Sohlbach, M., Weber, M. & Willhauck, G., 2001. eCognition Object Oriented Image Analysis. User Guide, Definiens Imaging, Munich. Ballinger, A. & Morgan, D. G., 2002. Validating two methods for monitoring population size of the

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European rabbit (Oryctolagus cuniculus). Wildlife Research, 29: 431–437. Beja, P., Pais, M. & Palma, L., 2007. Rabbit Oryctolagus cuniculus habitats in Mediterranean scrubland: the role of scrub structure and composition. Wildlife Biology, 13: 28–37. Benichou, P. & Le Breton, O., 1987. Prise en compte de la topographie pour la cartographie des champs pluviométriques statistiques. La Météorologie, 19: 23–34. Boag, B., 1987. Reduction in numbers of the wild rabbit (Oryctolagus cuniculus) due to changes in agricultural practices and land use. Crop Protection, 6: 347–351. Calvete, C., Estrada, R., Angulo, E. & Cabezas–Ruiz, S., 2004. Habitat factors related to wild rabbit conservation in an agricultural landscape. Landscape Ecology, 19: 531–542. Carvalho, J. C. & Gomes, P., 2004. Influence of herbaceous cover, shelter and land cover structure on wild rabbit abundance in NW Portugal. Acta Theriologica, 49: 63–74. Delibes, M. & Hiraldo, F., 1981. The rabbit as prey in the Iberian Mediterranean ecosystem. In: Proceedings of the I World Lagomorph Conference: 614–622 (K. Myers & C. D. MacInnes, Eds.). Univ. of Guelph, Canada. Delibes–Mateos, M., Redpath, S. M., Angulo, E., Ferreras, P. & Villafuerte, R., 2007. Rabbits as a keystone species in southern Europe. Biological Conservation, 137: 149–156. Hauck, W. W. & Donner, A., 1977. Wald’s test as applied to hypotheses in logit analysis. Journal of the American Statistical Association, 72: 851–853. Hosmer, D. W. & Lemeshow, S., 2000. Applied Logistic Regression, 2nd edn. John Wiley & Sons, Inc., New York. Jimenez, D., Martinez, J. E. & Peiro, V., 2006. Relationship between game species and landscape structure in the SE of Spain. Wildlife Biology in Practice, 2: 48–62. Lebreton, J. D., Burnham, K. P., Clobert, J. & Anderson, D. R., 1992. Modeling survival and testing biological hypotheses using marked animals: a unified approach with case studies. Ecological monographs, 62: 67–118. Lemeshow, S. & Hosmer, D. W., 1982. A review of goodness of fit statistics for use in the development of logistic regression models. American Journal of Epidemiology, 115: 92–106. Lombardi, L., Fernandez, N. & Moreno, S., 2007. Habitat use and spatial behaviour in the European rabbit in three Mediterranean environments. Basic and Applied Ecology, 8: 453–463. Lucas, R., Rowlands, A., Brown, A., Keyworth, S. & Bunting, P., 2007. Rule–based classification of multi–temporal satellite imagery for habitat and agricultural land cover mapping. ISPRS Journal of photogrammetry and remote sensing, 62: 165–185. Marchandeau, S., 2000. Enquête nationale sur les tableaux de chasse à tir saison 1998–1999. Le lapin de garenne. Faune Sauvage, 251: 18–25. Marchandeau, S., Chantal, J., Portejoie, Y., Barraud,

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S. & Chaval, Y., 1998. Impact of viral haemorrhagic disease on a wild population of European rabbits in France. Journal of Wildlife Diseases, 34: 429–435. Marchandeau, S., Devillard, S., Aubineau, J., Berger, F., Léonard, Y. & Roobrouck, A., 2007. Domaine vital chez le lapin de garenne dans trois populations contrastées. Rapport scientifique, ONCFS: 33–37. McGarigal, K. & Marks, B. J., 1994. Fragstats: spatial pattern analysis program for quantifying landscape structure (version 2.0). Forest Science Dept., Oregon State Univ. Moreno, S. & Villafuerte, R., 1995. Traditional management of scrubland for the conservation of rabbits Oryctolagus cuniculus and their predators in Doñana National Park, Spain. Biological Conservation, 73: 81–85. Moreno, S., Villafuerte, R. & Delibes, M., 1996. Cover is safe during the day but dangerous at night: the use of vegetation by European wild rabbits. Canadian Journal of Zoology, 76: 1656–1660. Palomares, F. & Delibes, M., 1997. Predation upon European rabbits and their use of open and closed patches in Mediterranean habitats. Oikos, 80: 407–410. Posada, D. & Buckley, T. R., 2004. Model Selection and Model Averaging in Phylogenetics: Advantages of Akaike Information Criterion and Bayesian Approaches Over Likelihood Ratio Tests. Systematic Biology, 53: 793–808. R Development Core Team, 2011. R: A Language and Environment for Statistical Computing. Available at: http://www.R–project.org/ [accessed 28 IX 11]. Rogers, P. M. & Myers, K., 1979. Ecology of the

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European wild rabbit, Oryctolagus cuniculus (L.) in Mediterranean habitats. I. Distribution in the landscape of Coto Doñana, S. Spain. Journal of Applied Ecology, 16: 691–703. Serrano Pérez, S., Jacksic, D., Meriggi, A. & Vidus Rosin, A., 2008. Density and habitat use by the European wild rabbit (Oryctolagus cuniculus) in an agricultural area of northern Italy. Hystrix, The Italian Journal of Mammalogy, 19: 143–156. Traissac, P., Martin–prével, Y., Delpeuch, F. & Maire, B., 1999. Régression logistique vs autres modèles linéaires généralisés pour l’estimation de rapports de prévalences. Revue d’Epidémiologie et de Santé Publique, 47: 593–604. Trout, R. C., Ross, J., Tittensor, A. M. & Fox, A. P., 1992. The effect on a British wild rabbit population (Oryctolagus cuniculus) of manipulating myxomatosis. Journal of Applied Ecology, 29: 679–686. Union européenne–SOeS, 2006. Corine Land Cover. Villafuerte, R., Calvete, C., Blanco, J. C. & Lucientes, J., 1995. Incidence of viral hemorrhagic disease in wild rabbit populations in Spain. Mammalia, 59: 651–659. Villafuerte, R. & Moreno, S., 1997. Predation risk, cover type and group size in European rabbits in Doñana (SW Spain). Acta Theriologica, 42: 225–230. Xiaoxia, S., Jixian, Z. & Zhengjun, L., 2005. A comparison of object–oriented and pixel–based classification approaches using QuickBird Imagery. In: ISPRS STM 2005 (ISSTM): Proceedings of International Symposium on Spatio–temporal Modeling, Spatial Reasoning, Analysis, Data Mining and Data Fusion. Peking Univ., China.

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Aerial ungulate surveys with a combination of infrared and high–resolution natural colour images U. Franke, B. Goll, U. Hohmann & M. Heurich

Franke, U., Goll, B., Hohmann, U. & Heurich, M., 2012. Aerial ungulate surveys with a combination of infrared and high–resolution natural colour images. Animal Biodiversity and Conservation, 35.2: 285–293. Abstract Aerial ungulate surveys with a combination of infrared and high–resolution natural colour images.— Information on animal population sizes is crucial for wildlife management. In aerial surveys, we used a silent light aircraft (microlight) and a combination of a computer–linked thermal infrared camera (640 x 480 pixels) to detect ungulates and high–resolution visual images (5,616 x 3,744 pixels) to identify specific species. From winter 2008/2009 to winter 2010/2011, we flew 48 missions over three German national parks and a German/ French biosphere reserve. Within each study area, we followed non–overlapping linear transects with a flying altitude ~450 m above ground level and scanned 1,500–2,000 ha every two hours of flight time. Animals best detected and identified were red deer and fallow deer. Detection rates with respect to the type and density of vegetation cover ranged from 0% (young spruce) to 75% (young defoliated beech) to 100% (open land). This non–invasive method is cost–effective and suitable for many landscapes. Key words: Aerial survey, Infrared camera, Microlight aircraft, Ungulates, Wildlife monitoring. Resumen Estudios aéreos de ungulados mediante una combinación de imágenes infrarrojas y naturales en color y alta resolución.— La información sobre el tamaño de las poblaciones animales es crucial para la gestión de la fauna salvaje. En los estudios aéreos, utilizamos un avión ligero y silencioso (ultraligero) y una combinación de una cámara infrarroja térmica (640 x 480 píxeles) conectada a un ordenador para detectar a los ungulados, e imágenes visuales de alta resolución (5.616 x 3.744 píxeles) para identificar las especies. Desde el invierno 2008/2009 al invierno 2010/2011 volamos en 48 misiones sobre tres parques nacionales alemanes y una reserva de la biosfera franco–alemana. En cada área de estudio, recorrimos transectos lineales no solapados con una altitud de vuelo aproximada de 450 m sobre el nivel del suelo, y escaneamos 1.500–2.000 ha cada dos horas de vuelo. Los animales que mejor se detectaron e identificaron fueron el ciervo común y el gamo europeo. Las tasas de detección con respecto al tipo y densidad de la cubierta vegetal fueron del 0% (píceas jóvenes), pasando por el 75% (hayas defoliadas jóvenes) al 100% (terreno abierto). Este método no invasivo tiene unos costes ajustados y es adecuado para muchos tipos de paisajes. Palabras clave: Estudio aéreo, Cámara infrarroja, Avión ultraligero, Ungulados, Estudio de la fauna salvaje. Received: 15 I 12; Conditional acceptance: 6 VII 12; Final acceptance: 31 VIII 12 Ulrich Franke & Bianca Goll1Aerosense Engineering, Auf dem Gries 1, 67280 Quirnheim, Germany.– Ulf Hohmann Inst. for Forest Ecology and Forestry, Division Forest and Wildlife Ecology, Hauptstr. 16, 67705 Trippstadt, Germany.– Marco Heurich, Bavarian Forest National Park, Dept. for Research and Documentation, Freyunger Str. 2, 94481 Grafenau, Germany. Corresponding author: Ulrich Franke. E–mail: u.franke@aerosense.de

ISSN: 1578–665X

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Introduction Data on the population size and/or density of larger mammals, such as ungulates, is essential for wildlife management, forestry, wildlife conservation, and land–use development (Apollonio et al., 2010). Several methods can be used to determine population densities of red deer (Cervus elaphus), fallow deer (Dama dama), roe deer (Capreolus capreolus) and wild boar (Sus scrofa), such as pellet counts, track surveys, spotlight counts and genetic capture recapture surveys (Tottewitz et al., 1996; D'Eon, 2001; Campbell et al., 2004; Garel et al., 2010; Ebert et al., 2012). However, in densely forested areas, there is as yet no accurate, robust and cost–effective method to count ungulates in a practical way (Strandgaard, 1972; Pielowski, 1984; Gaillard et al., 1998; Smart et al., 2004; Borkowski et al., 2011). Coordinated counts and drive counts are not very reliable and staffing costs are high. Track counts are restricted to a period with snow cover. Spotlight counts depend on a road network and good visibility in the forest stands (Focardi et al., 2001), Capture–Mark–Recapture studies are labour intensive and the laboratory work of genetic studies is currently still expensive. However, both methods could deliver accurate and precise results (Gill et al., 1996; Cederlund et al., 1998; Lukacs & Burnham, 2005; Curtis et al., 2009). Although aerial infrared surveys have been conducted in the past (Garner et al., 1995; Naugle et al., 1996; Haroldson et al., 2003; Bernatas & Nelson, 2004), they are not yet widely used to monitor animals, probably because of the relatively high costs and the lack of species–specific identification using only infrared images. Until now, in most aerial surveys, the camera operator actively searches for ungulates (Havens & Sharp, 1998; Bernatas & Nelson, 2004; Potvin & Breton, 2005). The disadvantage of this method is that the results depend heavily on the experience of the operator (Garner et al., 1995; Naugle et al., 1996; Haroldson et al., 2003). To date, mainly helicopters or twin– or single– engine aircrafts have been used. Helicopters, in particular, have the disadvantage that they potentially disturb wildlife, leading to biases in data aquisition. In 2004, we began to test thermal infrared cameras mounted on a silent and cost–effective microlight aircraft, which under German regulations is a light airplane with a maximum takeoff weight (MTOW) of 472.5 kg. In 2008 we began to improve and test the methodology. Our aim was to develop and test a standardized, non–invasive, observer–independent, and cost–effective method to count large mammals, especially in forested areas. Here we present the first results obtained while establishing the system. Material and methods Study sites Starting in 2004, we flew more than 60 missions over various types of areas, from flat land to mountainous regions and from land with open shrub cover to 100%

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forested regions with 60% coniferous trees. From winter 2008/2009 to winter 2010/2011, we flew 48 missions over three German national parks and a German/French biosphere reserve. The Bayerischer Wald National Park in south–eastern Germany is 24,250 ha with altitudes ranging from 600 to 1,450 m a.s.l. The vegetation consists mainly of high montane Norway spruce (Picea abies) forests and mixed mountain forest with European beech (Fagus sylvatica), Norway spruce, and white fir (Abies alba). The land cover of the southern part of the park is 16% coniferous forest, 25% deciduous forest, 26% mixed forest, 30% dead–wood stands as a consequence of an ongoing bark beetle (Ips typographus) outbreak (Lausch et al., 2010) and 3% open land. The northern part is covered with 39% coniferous forest, 19% deciduous forest, 31% mixed forest, 10% dead wood, and 1% open land. The following ungulate species occur in the park: red deer, roe deer, and wild boar (Heurich et al., 2011). The Hainich National Park is 7,500 ha and lies in central Germany. It consists of 51% deciduous forest, 30% open land, 16% pioneer forests, and 3% coniferous forest. Lying at around 450 m a.s.l., the park has only small altitude differences. It is inhabited by red deer, fallow deer, roe deer, and wild boar. The Kellerwald–Edersee National Park is 5,700 ha and is situated in central Germany. Altitudes range from 200 to 626 m a.s.l. It consists of 78% deciduous forest, 15% coniferous forest, 4% open land, and 3% undefined landscape. It is inhabited by red deer, fallow deer, roe deer, and wild boar. The central part of the German/French biosphere reserve Pfälzerwald–Vosges du Nord contains a 10,400 ha wildlife investigation area in south–western Germany. Altitudes range from 210 to 609 m a.s.l., and the area consists of 61% mixed coniferous forest, 29% deciduous forest, 6% open land, and 4% undefined landscape. It is inhabited by red deer, roe deer, and wild boar. Technical equipment For the aerial surveys, we used a silent, slow–flying microlight airplane (S–Stol) that has very low operating costs (fig. 1). The cost of operating the aircraft without technical equipment is about 100 €/h, compared to approximately 1000 €/h for a helicopter. The microlight plane was equipped with two camera mounts that provided a nearly vertical view towards the ground and a computer–linked camera system consisting of a JENOPTIC® infrared camera (640 x 480 pixels) and a Canon 5D Mark 2® high– resolution RGB camera (5,616 x 3,744 pixels). The infrared camera used a microbolometer detector that is sensitive to wavelengths of 7.5–14 µm and resolves temperature differences of 0.08 Kelvin. The lens had a field of view (FOV) of 12 x 9°, thereby providing a 96 m wide swath when flying 450 m above ground level. Data acquisition Sampling design To guarantee a standardized and transparent method,

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Fig. 1. Microlight aircraft used for aerial wildlife monitoring. Fig. 1. Avión ultraligero utilizado para el estudio aéreo de la fauna salvaje.

we flew pre–designed, non–overlapping transects that cover each investigation area completely (see example in fig. 2). The transects were 250 m apart, thus having about 150 m between two transects not scanned by the cameras. In mountainous regions, we planned the transects along the contour lines to minimize large altitude differences. The aircraft flew at approximately 100 km/h, and the cameras scanned 1,500 to 2,000 ha of study sites of 5,000 to 10,000 ha within two hours of flight time. We divided larger areas like the Bayerischer Wald National Park into two areas to maintain sampling sizes of one–third to one–fifth of the investigation area. Non–overlapping transects reduced the chances of double counting. To reduce the chances of double counting further, we had a closer look at the species, group size, and behaviour when detection events were close to each other on neighbouring transects. As infrared radiation does not penetrate vegetation, we flew most missions after leaves had fallen, from November to April. We also flew some missions in summer in areas where there was less tree coverage. In addition to testing the effect of the time of the year and the ambient temperature on the recording, we tested the time of the day and external radiation. Data recording and interpretation Both camera systems were connected to an onboard computer. The infrared camera produced a film, and the data was stored on a hard disk in real time. The advantage of digital data is that the temperature level and range can be later changed as appropriate, which is not possible with older systems where the data is stored on video tape. As weather and radiation conditions can change frequently during one flight, there was then no need to calibrate the temperature level and range during the flight. The visual images were stored on a compact flash card. Whenever a visual image was taken, the file number was written in

the infrared data stream to enable an easier association of infrared and visual images. A global positioning system (GPS) enabled the pilot to fly the predefined transects and to store the actual flight path. These data were used to geo–tag the detection sites. The data were analysed on a computer with two screens, one showing the infrared film (fig. 3A), and the other showing the visual images (fig. 3B). Whenever something raised the attention of the observer in the infrared film, the visual image was used to verify the detection. This enabled differentiation, for example, between an animal lying on the ground and a recently vacated and still warm resting area. If the observation was uncertain, the detection was not counted. As the detected radiation of an animal and its surrounding can vary, it was necessary to change the temperature level and ranges frequently and to rewind sequences to double check. The data were stored for reinterpretation or interpretation by another person. Test of dependence of detection rates on the degree of vegetation cover We worked only with the number of animals actually counted as a minimum number; we did not use correcting factors, such as the detection probabilities. Nevertheless, we tested the dependence of detection rates on the degree of coverage. We used mid–sized dogs as test animals at 116 positions on test transects with different types of vegetation cover: old spruce with animal close to the trunk, old spruce with animal between two trees, old defoliated beech with animal close to the trunk, old defoliated beech with animal between two trees, young spruce, young defoliated beech, and thicket, and open land as a reference. At each position, the degree of coverage was calculated by transforming a photograph taken vertically from the

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Fig. 2. Flight path along transects over the approximately 6,000 ha study site of the Kellerwald–Edersee National Park. All 89 detection sites are shown with one or several animals per site. Fig. 2. Trayectoria de vuelo a lo largo de los transectos, sobre aproximadamente 6.000 ha de la zona de estudio en el Parque Nacional Kellerwald–Edersee. Se indican los 89 lugares de detección, con uno o varios animales por lugar.

ground towards the sky into a binary black–and–white picture. We defined the degree of coverage as the ratio of pixels that showed the vegetation to the number of all pixels. Vegetation cover ranged from 0% for open land to 97% for young spruce forest. For each position, the degree of vegetation cover and the detection by the IR camera (yes or no) were noted.

Results Equipment and flight parameters To determine the footprint of the cameras and their geometrical resolution, we tested flying at altitudes of 300, 450, and 600 m above ground level over an enclosure with red deer. An altitude of 450 m provided

Fig. 3. Detection of animals using infrared camera (A) and close up of visual image taken in parallel to the infrared image, showing three male red deer (B). Fig. 3. Detección de animales utilizando la cámara infrarroja (A) y ampliación de la imagen visual tomada en paralelo con la imagen infrarroja, mostrando a tres ciervos comunes machos (B).

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the best combination of the size of the scanned area and geometrical resolution. At this flight height we did not observe any disturbance to wildlife caused by the aircraft either during our flights across enclosures or in any of our video recordings during our survey flights. Time of year/day and meteorological conditions We hypothesized that cold winter days with a snow cover would be the best time of the year to use thermal infrared since the temperature difference (ΔT) between the animal and the background should be at its highest. Unexpectedly, we observed, for example, small ΔTs on a winter flight at –13°C with snow–covered ground and large ΔTs on a summer flight with ambient temperatures of approximately 20°C. The ambient temperatures, therefore, were a weak factor at most. We tested the effect of the time of day on the recordings by flying from early morning to late evening. The best results were obtained around midday. One of the strongest influencing factors was the external radiation from sunlight. In bright sunlight, other objects, e.g. tree trunks and rocks, were sometimes misinterpreted as animals, and the emission of treetops sometimes prevented the detection of an animal standing underneath them. The best conditions found for the thermal infrared survey during the day was an 8/8 cloud cover (overcast).

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Table 1. Species of animals clearly identified using infrared and visual cameras mounted on a microlight aircraft. The data are from 28 flights flown from II 09 to IV 11: Na. Number of animals; Nd. Number of detections; Na/d. Number of animals per detection. Table 1. Especies de animales claramente identificados utilizando cámaras infrarrojas y visuales montadas en un avión ultraligero. Los datos proceden de 28 vuelos del II 09 al IV 11: Na. Número de animales; Nd. Número de detecciones; Na/d. Número de animales por detección.

Species

Na/d

Red deer

495

150

3.3

Fallow deer

249

4.8

Wild boar

239

3.0

Roe deer

1.3

Foxes

1.1

Wolves

3.0

Badgers

1.0

1,022

311

3.3

Total animals

Species detection and identification The success rate of verifying the thermal images with the visual, natural colour images increased from 60% in 2008/2009 to over 90% in 2010/2011. In most study sites, different species were present. Therefore, a clear distinction between the species was necessary. Infrared data alone did not suffice to distinguish between, for example, red deer and fallow deer in one study site, but with the visual imagery in addition, we were able to identify nine different species (table1). The species of animals detected was identified in 50% of cases. Animals in groups, e.g., red deer or fallow deer, were easier to detect than single animals. It was also easier to identify the species of a group, assuming that all animals in the group belonged to the same species, than to identify the species of a single animal, since the chance of having one animal clearly visible in the image increased with the number of animals in the group. Although the infrared signature of wild boar was also easy to distinguish, it was difficult to count exact numbers of animals because they often clustered close together. In all our surveys, roe deer were obviously underestimated, as we counted only 0 to 4 individuals per flight and harvest rates indicated much higher numbers. Successive flights and representativeness of samples We surveyed each of the four areas three times within as short a period as possible and every year for

three years, originally to obtain a statistical repetition. We soon realized that many factors influenced the results. The main factor influencing the aerial infrared surveys was the weather, which could allow us to fly two missions in one day or force us to wait a week between flights. Therefore, it was difficult to treat three successive flights as a statistical repetition, as the weather was seldom the same over several days. The repetitions were thereafter used to obtain one survey with the highest numbers only. Here we present the results of two examples of three successive flights over two different areas (table 2). The three flights over the Kellerwald–Edersee National Park were conducted during similar weather conditions (overcast, 12°C). The minimum population densities of red deer and fallow deer ranged from 5.5 to 6.6 animals/100 ha. With a coefficient of variation (CV) of 10%, the three flights yielded similar results. The three flights over the Hainich National Park were conducted during different weather conditions. There were overcast skies and ambient temperatures of 7°C during the first two flights and a 3/8 becoming a 5/8 cloud cover during the third flight with ambient temperatures of 13°C. The minimum population density of red deer and fallow deer was 4.4 animals/100 ha during the first two flights and only 1.3 animals/100 ha during the third flight. The CV was 52%. To check our samples for representativeness, we used statistics for one typical survey flight of the Hainich National Park, where we estimated a minimal

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Table 2. Comparison of the variation of data from three flights, each for two investigation areas: Min. Minimum number of red deer and fallow deer; DN(100). Population density (number/100 ha) (Mean, standard deviation [SD] and coefficient of variation [CV]). Table 2. Comparación de la variación de datos de tres vuelos, cada uno para dos áreas de investigación: Min. Número mínimo de ciervos común y gamo europeo; DN(100). Densidad de población/100 ha (Media, desviación estándard [SD] y coeficiente de variación [CV]). Site

Date

Min.

Area scanned

16 III 2009

1,074

4.4

18 III 2009

1,673

4.4

30 III 2009

1,709

1.3

09 IV 2010

111

1,680

6.6

10 IV 2010

1,768

5.5

14 IV 2010

103

1,768

5.8

Hainich

Kellerwald

DN(100)

DN(100) Mean SD CV[%] 3.4

1.8 52%

0.6 10%

density for all ungulates of 13.9 animals per 100 ha. From the flight data, we evaluated ten random subsamples of each sample size (P80, 80% of the whole flight; P70, 70% of the whole flight; P60, 60% of the whole flight, and P50, 50% of the whole flight) (fig. 4). The largest variation of 10.6 animals/100 ha to 13.9 animals/100 ha (23.7%) was found as expected within the P50 plot. The ten subsamples did not differ significantly, which is the first evidence that for this particular survey, the flight work could be reduced without losing representativeness. Test of dependence of detection rates on the degree of vegetation cover As expected, the detection rates increased with a decrease in forest cover (table 3). The detection rates ranged from 0% (young spruce) to 75% (young defoliated beech) to 100% (open land). We divided positions in old spruce and old beech forests into two subclasses ―with the test animal close to the tree trunk and with the test animal between two trees― to check whether the detection of an animal is blocked by a large tree trunk or by the many branches of neighbouring trees. In beech forests, the detection rate was best when the animal was located close to the trunk; in spruce forests, the detection rate was highest when the animal was located between two trees. Discussion The microlight airplane proved to be a good camera platform for wildlife surveys. The fact that the animals were not disturbed by the aircraft can be explained by the very low noise level of 59.1 db(A) of the used aircraft (noise certificate). In comparison, the noise level of an EC 135 helicopter we used for night flights

some years ago 84 db(A) was marked on the noise certificate. Moreover, the microlight plane can use a great variety of airstrips, including very small ones; it can be operated freely, and transfer times can be kept very short. In some countries, microlight aircrafts are allowed to land in any field with the consent of the owner. Costs are kept to a minimum because of the low fuel consumption. In the aerial surveys of Bernatas & Nelson (2004) and Haroldson et al. (2003), the flight pattern was determined by active searches for animals. This method has the advantages of large areas being covered and probably higher detection rates. Depending on the camera system, an active search also enables the operator to change to near field of view (NFOV) to identify the species with the IR camera only, thus also enabling night surveys with species–specific identification; however, only a small area can be surveyed while in the NFOV mode. The results of Haroldson et al. (2003) were highly variable; the sensor operation was inconsistent, the two operators had difficulty panning the cameras systematically, and the thermal contrast was variable. In an attempt to achieve a higher degree of standardization, we instead decided to use a predefined flight route with fixed cameras. The advantages of our standardized transect flight pattern were that there was no bias between different operators during the data acquisition and that the digitally stored data allowed multiple evaluations by different interpreters, thus making the method more transparent. Daniels (2006) assumed that fresh snow, complete cloud cover and low temperatures were optimal conditions for the IR technology, whereas our results indicated that the thermal IR technology can be used throughout the year in terms of temperature differences between animal and background. This is in agreement with Bernatas & Nelson (2004), who found that the

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291

Minimal density of ungulate (number/100 ha)

P70

P80

14 13 12

P50

P60

10 9 8 7

P50

6 5

P60

4 3

P70

P80

1 0

5 6 Subsample

Fig. 4. Representativeness of survey. Ten random subsamples of sample sizes of 80% (P80), 70% (P70), 60% (P60), and 50% (P50) of one entire survey flight over Hainich National Park were chosen, and the minimal density of all ungulates for each subsample of each sample size was calculated. Fig. 4. Representatividad del estudio. Se eligieron diez submuestras tomadas al azar con tamaños del 80% (P80), 70% (P70), 60% (P60) y 50% (P50) de un vuelo de estudio sobre el Parque Nacional Hainich, y se calculó la densidad mínima de todos los ungulados para cada submuestra de cada tamaño de muestra.

ambient temperature does not greatly influence the detection of wildlife with IR. During summer, foliage could interfere with thermal IR, but the technology can still be used if there is not much vegetation cover. In forested regions, it is necessary to conduct surveys during the months when the trees are defoliated. Aerial surveys with thermal IR at night time have the advantage of minimal jamming by other radiation. During the day, other objects, such as rocks and tree trunks might become heated or reflect radiation and they can then be misinterpreted as a detection event. However, we had to fly between sunrise and sunset, first because microlight aircraft in Europe are only allowed to be operated during daytime, and second, because we used visual images for identification. The best results were obtained around midday. This finding is in accordance with the results of Arnold et al. (2004), who described that the subcutaneous temperature of red deer in winter could be more than 10°C lower in the early morning than at other times. But in our opinion, the most important factor for infrared surveys during the day are homogeneous conditions, preferably with an 8/8 cloud cover. With higher numbers of visual images recorded during one flight, we were able to increase species’

identification to about 50% of all detected animals. Although we only used the visual data to identify the species as these data are much clearer than IR data, if we were to use the different IR signatures of the species, we could achieve a higher percentage of species identification. If only species living in groups (e.g. red deer or fallow deer) are of interest, species identification would be greater than 50% because at least one animal can usually be identified in the visual image. Roe deer were underestimated in our studies, most likely because of their solitary mode of life in forest habitats, their smaller body size, and their ability to hide in thick vegetation. We detected many smaller IR signatures, but could not identify them on the visual images. In tests with roe deer in fields and with roe deer that we positively identified, we obtained thermal IR signals; the lack of thermal radiation is not therefore the reason for their underestimation. If we disregarded all the single detection events of small animals and objects, the rate of species identification would be much higher. Our comparisons of three consecutive flights indicated that the data cannot always be used as a statistical repetition. The data from the Hainich National Park survey showed a large variance. We think that

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Table 3. Dependence of detection rates on the degree of vegetation cover. Test animals were placed at the indicated positions, and their detection by infrared and visual cameras was tested. Tabla 3. Dependencia de las tasas de detección del grado de la cubierta vegetal. Los animales de prueba se situaron en las posiciones indicadas, y se estudió su detección mediante cámaras infrarrojas y visuales. Vegetation n Old defoliated beech, animal close to trunk

Average degree of cover (%)

Detection rate (%) 94

Old defoliated beech, animal between two trees

Young defoliated beech

Old spruce, animal close to trunk

Old spruce, animal between two trees

Young spruce

Thicket

Open land

100

this is not due to the 12 days between the second and the third flight, but rather to worse weather conditions on 30 III 09. This notion is supported by the fact that the weather conditions during the three flights over the Kellerwald–Edersee National Park were similar and the results varied only little. Our statistical test of subsamples of the survey showed that our sample size was representative, at least when all ungulates were considered together. Further tests of species often present in larger groups, e.g. red deer or fallow deer, are needed. In our test of dependence of detection rates on the degree of vegetation cover, the interpreter of the infrared data was aware that dogs were present as test animals. In addition, the dogs were placed more or less on the centre line of the transect to avoid animals being outside the camera range. This knowledge and set up probably led to higher detection rates. In future evaluations, we intend to mix the images and include some showing no animals. The use of airborne thermal and visual imagery to monitor wildlife has the advantage of sampling larger areas per time unit than other methods. Furthermore, there is no dependence on roads and tracks. The combination of a thermal infrared and a visual camera together with the stored data makes the method even more transparent. We believe that this method will be used more frequently in the future. Acknowledgements We thank the Deutsche Bundesstiftung Umwelt, which funded the in–depth research during the last three years. The study was also funded by the Baviarian Forest National Park, the Hainich National Park, the Kellerwald–Edersee National Park, and the FAWF (Trippstadt), which is responsible for the wildlife inves-

tigation area of the biosphere reserve Pfälzerwald ― Vosges du Nord. We thank Felix Wilmes for organizing the test of the detection rates (thesis project), Prof. Dr. Roland Klein (Universität Trier) for help with statistics, Michael Hornschuh (Hainich National Park), Wolfgang Kommallein (Kellerwald–Edersee National Park), and last but not least, we would like to thank all the other people who helped in one way or another over the past seven years while we were working in this field. References Apollonio, M., Andersen, R. & Putman, R., 2010. European ungulates and their management in the 21st century. Cambridge Univ. Press. Arnold, W., Ruf, T., Reimoser, S., Tataruch, F., Onderscheka, K. & Schober, F., 2004. Nocturnal hypometabolism as an overwintering strategy of red deer (Cervus elaphus). Am. J. Physiol. Regul. Integr. Comp. Physiol., 286: R174–R181. Bernatas, S. & Nelson, L., 2004. Sightability model for California bighorn sheep in canyonlands using forward–looking infrared (FLIR). Wildlife Society Bulletin, 32: 638–647. Borkowski, J., Palmer, S. C. F. & Borowski, Z., 2011. Drive counts as a method of estimating ungulate density in forests: mission impossible? ActaTheriol., 56: 239–253. Campbell, D., Swanson, G. M. & Sales, J., 2004. Comparing the precision and costeffectiveness of faecal pellet group count methods. Journal of Applied Ecology, 41(6): 1185–1196. Cederlund, G., Bergqvist, J., Kjellander, P., Gill, R. B., Gaillard, J. M., Boisaubert, B., Ballon, P. & Duncan, P., 1998. Managing roe deer and their impact on the environment: maximizing the net benefits to society. In: The European Roe Deer: the Biology of

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Success: 337–372 (R. Andersen, P. Duncan & J. D. C. Linnell, Eds.). Scandinavian Univ. Press, Oslo. Curtis, P. D., Boldgiv, B., Mattison, P. M., & Boulanger, J. R., 2009. Estimating deer abundance in suburban areas with infrared–triggered cameras. Human–Wildlife Conflicts, 3(1): 116–128. Daniels, M. J., 2006. Estimating red deer Cerphus elaphus populations: an analysis of variation and cost–effectiveness of counting methods. Mammal Review, 36: 235–247. D’Eon, R. G., 2001. Using snow–track surveys to determine deer winter distribution and habitat. Wildlife Society Bulletin, 29: 879–887. Ebert, C., Knauer, F., Spielberger B., Thiele, B. & Hohmann, U., 2012. Estimating wild boar Sus scrofa population size using faecal DNA and capture–recapture modelling. Wildlife Biology, 18(2): 142–152. Focardi, S., De Marinis, A., Rizzotto, M. & Pucci, A., 2001. Comparative evaluation of thermal infrared imaging and spotlighting to survey wildlife. Wildlife Society Bulletin, 29(1): 133–139. Gaillard, J. M., Liberg, O., Anderson, R., Hewison, A. J.M . & Cederlund, G., 1998. Population dynamics of roe deer. In: The European roe deer: the biology of success. 309–333 (R. Anderson, P. Duncan & J. D. C. Linell, Eds.). Scandinavian Univ. Press, Oslo. Garel, M., Bonenfant, C., Hamann, J. L., Klein, F. & Gaillard, J. M., 2010. Are abundance indices derived from spotlight counts reliable to monitor red deer Cervus elaphus populations? Wildlife Biology, 16(1): 77–84. Garner, D. L., Underwood, H. B. & Porter, W. F., 1995. Use of Modern Infrared Thermography for Wildlife Population Surveys. Environmental Management, 19(2): 233–238. Gill, R. M. A., Johnson, A. L., Francis A., Hiscocks, K. & Peace. A. J.,1996. Changes in roe deer (Capreolus capreolus L.) population density in response to forest habitat succession. Forest ecology and management, 88(2): 31–41. Haroldson, B. S., Wiggers, E. P., Beringer, J., Hansen, L. P. & McAninch, J. B., 2003. Evaluation of aerial thermal imaging for detecting white–tailed deer in

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a deciduous forest environment. Wildlife Society Bulletin, 31: 1188–1197. Havens, K. J. & Sharp, E. J., 1998. Using thermal imagery in the aerial survey of animals. Wildlife Society Bulletin, 26(1): 17–23. Heurich, M., Baierl, F., Günther, S. & Sinner, K. F., 2011. Management and Conservation of large mammals in the Bavarian Forest National Park. Silva Gabreta, 17(1): 1–18. Lausch, A., Fahse, L., & Heurich, M., 2010. Factors of the spatial–temporal dispersion of bark beetle in the Bavarian Forest National Park from 1990 to 2007 – a quantitative landscape–level–analysis. Forest Ecology and Management, 261(2): 233–245. Lukacs, P. M. & Burnham, K. P., 2005. Review of capture–recapture methods applicable to noninvasive genetic sampling. Molecular Ecology, 14(13): 3909–3919. Naugle, D. E., Jenks, J. A. & Kernohan, B. J., 1996. Use of thermal infrared sensing to estimate density of white–tailed deer. Wildlife Society Bulletin 24(1): 37–43. Pielowski, Z., 1984. Some aspects of population structure and longevity of field roe deer. Acta Theriologica, 40: 197–217. Potvin, F. & Breton, L., 2005. From the field: Testing 2 aerial survey techniques on deer in fenced enclosures – visual double–counts and thermal infrared sensing. Wildlife Society Bulletin, 33(1): 317–325. Smart, J. C. R., Ward, A. I. & White P. C. L., 2004. Monitoring woodland deer populations in the UK: an imprecise science. Mammal Review, 34: 99–114. Strandgaard, H., 1972. The roe deer (Capreolus capreolus) population at Kalö and the factors regulating its size. Danish Review of Game Biology, 7: 1–205. Tottewitz, F., Stubbe C., Ahrens, M., Dobias, K., Goretzki, J & Paustian K. H., 1996. b Counting droppings as a method of estimating the population of ruminant game. Zeitschrift fürJagdwissenschaft, 42(2): 111–122.

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Common–interest community agreements on private lands provide opportunity and scale for wildlife management L. A. Powell

Powell, L. A., 2012. Common–interest community agreements on private lands provide opportunity and scale for wildlife management. Animal Biodiversity and Conservation, 35.2: 295–306. Abstract Common–interest community agreements on private lands provide opportunity and scale for wildlife management.— Private lands are critical to conservation planning for wildlife, worldwide. Agriculture subsidies, tax incentives, and conservation easements have been successfully used as tools to convert cropland to native vegetation. However, uncertain economies threaten the sustainability of these incentives. The wildlife management profession is in need of innovative models that support effective management of populations. I argue that biologists should consider the option of facilitating the development of private reserves to reduce the dependence of conservation on public investment. Private reserves can be enhanced by creating common– interest communities, which reduce the problem posed by limited size of individual properties. Cross–property agreements between landowners can provide economic incentives through forms of ecotourism, energy production, and/or enhanced agricultural production. I share two case studies that demonstrate how cross–property agreements may be beneficial to landowner’s finances and conservation of diverse wildlife communities, as well as providing an efficient structure for NGOs and management agencies to engage and support landowners. Key words: Conservation biology, Conservancy, Economics, Landscape, Policy, Private lands. Resumen Acuerdos comunitarios de interés común sobre los terrenos privados proporcionan oportunidades y extensión para la gestión de la naturaleza salvaje.— En todo el mundo, los terrenos privados son críticos para la planificación de la conservación de la naturaleza salvaje. Los subsidios agrícolas, los incentivos fiscales y las servidumbres para la conservación han sido utilizados con éxito como herramientas para convertir las tierras de cultivo en vegetación nativa. Sin embargo, las incertidumbres económicas amenazan la sostenibilidad de dichos incentivos. La gestión profesional de la naturaleza salvaje precisa de la innovación de los modelos que dan soporte efectivo a la gestión de las poblaciones. Opino que los biólogos deberían considerar la opción de facilitar el desarrollo de reservas privadas, con el fin de reducir la dependencia de la conservación basada en las inversiones públicas. Puede estimularse la creación de reservas privadas creando comunidades de interés común, que reduzcan el problema impuesto por el tamaño limitado de las propiedades individuales. Los acuerdos entre propietarios sobre sus terrenos pueden proporcionar incentivos económicos en forma de ecoturismo, producción de energía y/o una mejor producción agrícola. Comparto los estudios de dos casos que demuestran cómo los acuerdos entre propiedades pueden beneficiar tanto a las finanzas de los propietarios de las tierras como a la conservación de diversas comunidades silvestres, así como proporcionar una estructura eficaz para ONGs y agencias de gestión en el compromiso de dar soporte a los propietarios. Palabras clave: Biología de la conservación, Preservación, Economía, Paisaje, Política, Terrenos privados. Received: 13 II 12; Conditional acceptance: I VII 12; Final acceptance: 5 IX 12 Larkin A. Powell, 3310 Holdrege Street, 419 Hardin Hall, School of Natural Resources, Univ. of Nebraska–Lincoln, Lincoln, NE 68583–0974, USA. E–mail: lpowell3@unl.edu ISSN: 1578–665X

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Introduction Private lands are critical to conservation planning for wildlife worldwide. However, conservation efforts have primarily focused on public lands (Knight, 1999; Brown, 2010). Public lands certainly offer permanency of purpose, and public lands biologists enjoy a large degree of ownership of the decision to implement management decisions on the landscape. These strategies may be effectual when public lands dominate a region or country (fig. 1, table 1). But, private lands (defined as land under freehold or leasehold by individuals, not including land native communal lands; Swift et al., 2003) comprise the majority of many countries’ land base (table 1). Thirty–six of 50 US states have > 75% of their area managed by private landowners (fig. 1). Private lands biologists work in an arena in which they can only offer support for decisions in a landscape that is highly volatile with regard to alternative land uses. Both game species and threatened species stand to gain from well–positioned strategies for conservation on private lands. Conservation on private lands has emerged as a critical direction (Knight, 1999). Here, I describe challenges to private lands conservation from logistic and ecological perspectives. I suggest that common–interest communities should be considered as a viable option to create incentives for conservation on private reserves, while also providing scale that can support ecological processes that lead to successful conservation efforts. I assess the most common incentive tools for conservation on private lands. And, I provide two case studies to support Schultz’s (2010) suggestion that society’s demand for natural places may operate to encourage private landowners to work across property lines to produce goods and services that large, intact landscapes can provide. Challenges Challenges to provide incentives The primary challenge to conservation on private lands is to provide an incentive to landowners, as conservation measures may conflict with ventures designed to realize economic value from the land investment. Simply stated, landowners/investors must realize a profit. Aldo Leopold, writing in the 1940s, expressed frustration with farmers in Wisconsin who did not continue to implement soil conservation measures after an initial 5–year period of public investment of labor and machinery (Leopold, 1949). That frustration pervades the ranks of conservationists today, especially those who are not empathetic with the notion that resilient conservation practices must stand on the shoulders of economically resilient farm and ranch ventures. Wunder (2000) stated that the success of conservation incentives depends on the structure inherent in the mode of participation—how does conservation compare with other productive activities? Conservation incentives will have conservation impact only if they change labor and land allocation decisions on a

Powell

sustainable basis. In the US, the incentive challenge is perhaps greatest in regions with productive soils and adequate precipitation (fig. 2), where row–crop agriculture is the wise investment on the landscape because of record high prices for corn (fig. 3; July 2012 spot market, Nebraska USA: US$275–314/ metric ton), soybeans (US$588–624/metric ton), and wheat (US$293–330/metric ton). Agriculture subsidies The Food Security Act of 1985 ('Farm Bill'; Brown, 2010) served as a subsidy program to address concerns of soil erosion (wind and water) and price supports in the US. Conservation efforts are now implicit objectives in the current Farm Bill, and Farm Bill programs have been used as the primary method to convert cropland to native vegetation with successful short–term benefits (Haufler, 2005). Indeed, the job title of private lands biologists in the US is commonly 'Farm Bill Biologist'. The Farm Bill is an extensive government program; the Conservation Reserve Program (one program within the Farm Bill) paid US$1.7 billion in annual payments in 2012 to 737,699 contracts (most 10–year) on 409,253 farms (11,975,550 ha; United States Department of Agriculture, 2012). Similar subsidy programs are available in Europe through the Common Agricultural Policy (Pain & Pienkowski, 1997), but are not an option in most Latin American Countries because of budget priorities (Swift et al., 2003). It is very reasonable to expect that the combination of the US’ current budget shortfalls and the high rental rates now needed to compete with current rental rates offered for production purposes (fig. 3) may result in a loss of the diversity of types of direct payments in the next Farm Bill that would benefit wildlife habitat. There is no argument that subsidy programs have created benefits for wildlife on millions of acres in the US (Haufler, 2005), but the future of this program as the primary means to support conservation on private lands is in doubt. Tax incentives A recent development in several states in the US is the availability of tax incentives to farmers who pledge to keep their land in agricultural production or landowners with forests who pledge to manage them in an approved manner (Salkon et al., 2001). Agricultural tax credits may be useful to wildlife conservation in regions with a high degree of urban expansion, as the incentive may keep the land owner from transforming the farm into residential communities or industrial complexes. For example, in Nebraska, USA, a Greenbelt Tax was created to reduce development of urban areas along rural, riverine corridors (T. LaGrange, Nebraska Game and Parks Commission, personal communication). Agricultural landowners typically see their property values increase if urban growth creates development potential for their farm land. A person owning 65 ha in Lancaster County, Nebraska, for example, would be levied an additional US$2600 in annual property taxes if their land’s value increased from US$1200/ha to US$3600/ha. Such an increase might surpass current economic margins for crop

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0–50 51–75 76–90 91–95 96–100

Fig. 1. Percent of area under private ownership (non–state or non–federal) in the states of the USA, based on data collected by the Natural Resource Council of Maine. Fig. 1. Porcentaje de propiedad privada (ni estatal, ni federal) en los estados de EUA, basado en datos recogidos por el Consejo de Recursos Naturales de Maine.

production; it might be seen as especially unpalatable for a landowner interested in non–agricultural, recreational uses. A housing development or office complex could result. The Green Belt Tax status would allow the landowner to pay property taxes based on the agricultural value of the land rather than the full market value. It should be noted that the Greenbelt Tax only applies to agricultural and horticultural uses; a private nature reserve would not qualify for the incentive. However, the precedent of tax incentives exists, and state laws could be structured to provide tax incentives to private landowners who dedicate and properly manage a private nature reserve. Salkon et al. (2001) noted that tax incentives do not provide long term security for conservation efforts (table 2). Tax incentives for conservation have been used in some Latin American countries, but they are often withdrawn in times of economic insecurity (Swift et al., 2003). Even if the tax incentive remains, the price paid by developers for land may eventually exceed the maximum compensation through tax relief. Conservation easements Easements are used throughout the US (Salkon et al., 2001), Latin America (Swift et al., 2003), and Europe (conservation covenants; Kiesecker et al., 2007) by private landowners to restrict the future uses of their

property. A land trust is often created to be the recipient of the benefits of the easement (Schutz, 2010), and the recipient purchases the easement from the landowner. Easements are attractive to landowners, because the land remains in private ownership and owners continue to live on the land and derive benefit from farming, ranching, forestry, or other activities. Landowners often receive income from the sale of the easement, and this sale value varies by the market value of the land, the conservation need for the property (in fact, some landowners may find it difficult to find a third party with interest to purchase an easement), and local agreements. A landowner may donate a portion or all of the sale of the easement back to the land trust, reducing their income from the easement sale. However, their contribution may be considered a charitable donation, which can provide significant income and/or estate tax benefits derived from the state and federal government (Salton et al., 2001). Conservation easements are usually designed to be perpetual in nature (table 2). But, as Schutz (2010) noted, such easements are enabled by state legislation, and easements are the common subject of legislation (e.g., Legislature of Nebraska, 2012) that would affect their use. Of primary concern in rural districts is the potential loss of property tax, and thus support to county government and local schools. Easements are often the first step

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Table 1. Percentage of land area of select countries that is in private ownership: a Private ownership statistics not available; percentage represents percent of country in agricultural land use; b China practices public land ownership with no freeholds, only leaseholds. Tabla 1. Porcentaje de tierra de propiedad privada de los países seleccionados: a Estadísticas de propiedad privada no disponibles; el porcentaje representa el tanto por ciento de la tierra con uso agrícola; b China practica la propiedad pública de la tierra, que no permite la propiedad absoluta, sino únicamente el arrendamiento. Country

% Source

Australia

15 Forbes (1985)

Canada

10 Cahill & McMahon (2010)

China

0b Ho (2001)

Ethiopia

10a Cahill & McMahon (2010)

Germany

52a Cahill & McMahon (2010)

Latin American countries (most) > 80 Swift et al. (2003) Namibia

43 Shaw & Marker (2011)

Spain

83a Cahill & McMahon (2010)

Tanzania

11a Cahill & McMahon (2010)

United Kingdom > 80 Harrison et al. (1977) United States, excluding Alaska 75 NRCS (2001) Zimbabwe

42a Cahill & McMahon (2010)

towards the eventual sale of the property to a state or federal wildlife agency that may pay no property taxes, or may pay property taxes at a much lower property tax rate than a private landowner (Lancaster County, Nebraska, standard rate: US$0.0027/$100 valuation; Lower Platte South Natural Resources District rate: US$0.0004/$100 valuation; Lancaster County, 2011). The loss of property tax to Lancaster County, USA for the 65–ha property in the previous example, would be $3900/year for land worth US$3600/ha. Ten landowners making a similar decision would reduce income to the County equal to one government salaried worker (e.g., teacher, road maintenance, social aide). Therefore, the benefit to the individual (tax relief) is seen as a cost to the local community. Creation of reserves A governmental or non–governmental organization (NGO) may purchase a tract of land from an individual for the purposes of creating a nature reserve. The incentive to the individual is the fair market price (or

sometimes premium price) paid by the government or NGO at the time of the sale. Thus, the land is removed from development and can be restored to native vegetation or protected in a native state. Individuals can also develop private reserves on their land, although the official recognition (and economic incentives, if any) of private reserves varies from country to country, as well as from state to state within the US (Teer, 1999; Salkon et al., 2001; Swift et al., 2003). Private reserves are routinely considered in the set of tools available to wildlife biologists engaged with private landowners in Latin America (Swift et al., 2003) and southern Africa (Powell, 2010; Shaw & Marker, 2011), but they are not typically considered by biologists in the US (Salton et al., 2001) with the exception of the state of Texas (Teer, 1999). One factor in this shortcoming is that wildlife management students in the US are rarely required to take courses in business, tourism, or entrepreneurism, while their counterparts in nature conservation in southern Africa or Latin America (as examples) receive their education in the context of the economic benefits of properly managed populations of wildlife. Wildlife biologists have long been aware of the potential value from hunting, bird watching, nature walks, and environmental education on private lands, although resource ownership issues are complex (Freese, 1998; Teer, 1999; Thompson & Edwards, 2009). Such values are subject to variability in tourism markets, and the size and location of the reserve will affect its value to regional biodiversity, its draw to tourists or hunters, and its capacity to provide economic benefit to the land owner. These constraints or perceived risks may lead land owners to make the decision to sell their property to a government entity or NGO. Schutz (2010) suggests that private reserves could be supported by government during initial development to reduce these risks. The advantage to private reserves is that they are not usually dependent on public subsidies, and the reserves generate profits as a private venture (table 2). Private land conservation has typically concentrated on methods that have substantial cost in public investment (through purchase to create a public reserve or payment of annual subsidy), as well as eventual loss of agriculture productivity and contribution to taxes (Salton et al., 2001). Challenges to support landowner decisions Wildlife management decisions are complex and have a level of uncertainty, even when made by trained wildlife biologists. Thus, private landowners face the same challenge of making smart decisions, and should be trained in decision–making processes that include the need for clarifying objectives, assessing alternative management options, assessing potential risk of alternatives, and coordinating decisions with other current decisions (Gregory & Keeney, 2002). Monitoring to determine the level of success of a management decision is also critical (Lyons et al., 2008). Landowners are not typically trained in concepts or techniques of wildlife management or conservation biology. Fortunately, farmers and ranchers are usually trained to manage domestic plants and animals

Relative economic return/ha

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Row crops Cattle Wildlife

A B Rainfall/soil productivity

Fig. 2. Relative economic return (per ha) of three potential commodities across a gradient of rainfall and soil productivity from arid climate (left) to more mesic climate (right). A describes thresholds where cattle grazing becomes more profitable than wildlife–based entrepreneurial activities, and B is threshold where row crops become more profitable than cattle grazing. Lines showing relative economic return would be expected to shift with market conditions. Fig. 2. Retorno económico relativo (por ha) de tres productos potenciales a través de un gradiente de precipitación y producción del suelo desde un clima árido (izquierda) a un clima más moderado (derecha). A describe los umbrales donde el apacentamiento del ganado se hace más aprovechable que las actividades emprendedoras basadas en la fauna salvaje, y B es el umbral donde las cosechas se hacen más provechosas que el pastoreo del ganado. Sería de esperar que la línea que muestra el retorno económico relativo cambiase junto las condiciones del mercado.

(Powell, 2010), so concepts of population growth, competition, and sustainable harvest are familiar. Governments, agencies, NGOs, and universities have a critical role to provide for education of landowners (Swift et al., 2001). Training needs may be significant, which will result in costs to the agencies or NGOs. Ecological challenges to private lands conservation Swift et al. (2001) suggested that there are implicit ecological challenges to addressing conservation concerns on private lands: (1) size limitations of private property, (2) ad hoc locations of reserves in relation to priority conservation areas, and (3) the need for long–term sustainability of a conservation system (table 2). Property size limitations Conservation of biodiversity necessitates a diverse set of habitats (Toombs et al., 2010). The potential heterogeneity of habitats on a parcel of land increases with the size of the property. Farm– or ranch–level heterogeneity can be expected to be lower than landscape–level heterogeneity, because of farm– and ranch–level management decisions (e.g., type of grazing system or crop selection). As such, it would be rare for a single property to provide the diverse array of habitats needed for the conservation of a diverse community. So, biologists must engage with multiple landowners across

the landscape to achieve most conservation goals in traditional incentive programs (table 2). Second, the annual home range of most species of wildlife goes beyond the borders of a single property (e.g., sage grouse [Centrocercus urophasianus] mean annual movements: 11.3 km: Connelly et al., 1988; typical movements of > 10 km for white–tailed deer [Odocoileus virginianus] and mule deer [Odocoileus hemionus]: Frost et al., 2009). A landowner’s efforts to support the breeding needs of a deer population, for example, could be thwarted by a neighboring landowner’s overharvest during the fall. Efficient and effective use of conservation funds necessitates that the scale of animal movements be contained within the scale of conservation efforts (Scott et al., 1999). Last, private reserves that use iconic species for hunting or non–consumptive income face the challenge that many of these species occur at relatively low densities (Freese, 1998). Sustainable trophy harvest of white–tailed deer (Odocoileus virginianus), for example, requires that hunters follow a strategy for take that allows deer to grow older and reach trophy status, as judged by antler size (Jenks et al., 2002). A single landowner, by the merits of the number of trophy deer required for profitable operation (one multi–day hunt for a trophy deer, including meals, guiding, and lodging may be approximately US$5,000), would have to own thousands of acres to engage in a sustainable venture.

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Table 2. Comparison of selected incentive programs for conservation on private lands, with respect to common ecological challenges after Swift et al. (2001). Programs are categorized with regard to the source of the motivation to meet the challenge: Landowner. Challenge overcome through internal landowner motivations; Public assistance. Challenge overcome through external motivations from government or NGOs; No. Challenge not likely to be overcome; a Challenge met only with considerable effort to target several neighbors; b Challenge met only through efforts to provide higher incentive to landowners in a selected watershed or region. Tabla 2. Comparación de los programas de incentivación para la conservación de los terrenos privados, con respecto a los desafíos ecológicos comunes según Swift et al. (2001). Los programas están clasificados según la motivación para enfrentarse al desafío: Landlowner. Propietario, enfrentarse al desafío por las motivaciones internas del propietario; Public Assistance. Asistencia pública, enfrentarse al desafío por motivaciones externas del gobierno o las ONGs; No. No es probable que se enfrente al desafío; a Desafío encarado únicamente con considerable esfuerzo para incluir a varios vecinos. b Desafío encarado únicamente a través de los esfuerzos para proporcionar mayores incentivos a los propietarios de la tierra en una región o cuenca hidrológica determinada.

Ecological challenge

Incentive Limited size

Protects priority location

Long–term sustainability

Agriculture subsidies

Public assistance a

Public assistance b

Tax incentives

Public assistance

Conservation easements

Public assistance

Landowner

Private reserve: single owner

a a

b b

Private reserve: common–interest community Landowner No Landowner

Ad hoc location of private lands Conservation biologists often identify ‘gaps’ in the landscape that are not protected by public reserves, yet are critical to a species of conservation concern (Scott et al., 1993). Private reserves have the potential to fill such gaps, but not all private properties are positioned to connect corridors or create buffers around public areas (Swift et al., 2001) and thus complete a conservation strategy. Regardless of the tool used to provide incentive for conservation, this challenge will continue to require wildlife biologists to prioritize the geographic scope of their efforts on private lands (table 2). The need for long–term protection Conservation strategies should be aimed to increase resilience. Humans have reduced the resilience of agroecosystems by removing diversity and altering disturbance regimes. As altered systems, they may be more vulnerable to perturbation, and may quickly shift from a desired to less desired state (Folke et al., 2004). The perturbation may be ecological in nature (e.g., drought), but also political, social, or economic. The conservation incentives offered by agriculture subsidies, while affecting dramatic acreage of land (Barbarika, 2009), are not resilient to economic fluctuations. Grain prices (e.g., maize; fig. 3) are highly unpredictable

from year, which creates instability for long–term conservation because tradeoffs between subsidy payments and potential income from crop production are in constant flux. The benefits of local conservation efforts (e.g., Negus et al., 2010; Matthews, 2009) can disappear when incentives become less attractive than another investment option (fig. 4). Of the incentives traditionally used by private lands biologists, only conservation easements allow for long–term landscape transformation with the assumption that enabling legislation is not withdrawn. In contrast, owners of single– and multiple–owner private reserves have internal incentives to be successful over long periods of time, because of their personal investments in their ventures (table 2). Common interest communities Potential Private landowners who are interested in innovative, entrepreneurial conservation efforts will often have a need to work beyond the property limits of their land. Schutz (2010) suggested that common–interest communities may be a viable means of distributing benefits from nature–based entrepreneurial efforts on landscapes. A common–interest community is defined as an association of willing participants who accept

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Annual price 5–year average

US$ per bushel

4 3 2 1 0 1900

1920

1940

1960 Year

1980

2000

2020

Fig. 3. Annual market price paid for corn (US$ per bushel) in Nebraska, USA during 1908–2011. The 5–year moving average is shown as a dotted line (data from the National Agricultural Statistics Service, USDA). Fig. 3. Precio de mercado pagado anualmente por el maíz (US$ por cada 52 libras) en Nebraska, USA, durante el periodo 1908–2011. La línea de puntos muestra el promedio de la variación (datos del Servicio Estadístico Nacional de Agricultura, USDA).

rights and duties that are inherent in title to their real estate (Schutz, 2010). A common example is a homeowners’ association found in urban development. Schutz (2010) argues that the common–interest community model could easily be extended to include private lands for the benefit of wildlife populations. Owners of a parcel of land within a present–day lake association, for example, are obligated to engage in and/or refrain from certain uses of their land. The association might hold a lake as association property for the benefit of the owners in common. Management of the fishery is an example of services performed by the association for its members, who may be regulated on the type of dock or boat housing they may construct with an eye toward holding property values at high levels for all members (Korth & Klessig, 1990). Another form of common–interest community is a timber cooperative (Barten, 2001). Small, private landowners form agreements to market timber as an association to derive higher prices. The land remains in private ownership, but decisions on timber harvest are made as a group. As forest management is an indirect form of wildlife management, timber cooperatives are well–suited to develop additional income streams such as hunt leases or hiking retreats. It is easy to imagine the formation of a common–interest community by neighboring farmers or ranchers. Such arrangements between neighbors can provide participants with geographically larger operations and greater economic return without purchasing more land,

while also providing the legal framework in which to make joint decisions and to distribute costs and income among the participants. As such, common–interest communities would be well–suited to be used by private landowners with interests in creating a private reserve to support nature–based, entrepreneurial ventures. Benefits of scale for wildlife The formation of a common–interest community among neighbors results in the joint management of parcels of land. The co–managed landscape could be suitable for effective management of wildlife. This larger landscape under management allows structural heterogeneity of habitat to be established at multiple scales (Toombs et al., 2010), which further support diverse communities and protect rare species (Naidoo et al., 2011). In contrast, subsidy programs, tax incentives, conservation easements, and single–owner private reserves cannot, per se, provide the scale needed for conservation (table 2). Large, co–managed properties allow the establishment of 'zones' for management. Zones might be constructed around habitat types. Larger reserves allow more zones for different activities; more habitat zones should also result in more species of wildlife (Toombs et al., 2010), facilitating diverse use by tourists and increasing economic return (Naidoo et al., 2011). Zonation can also be used to set aside portions of the reserve for specific uses. For example, four

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Acres under contract in NE (USA)

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1,600,000 1,400,000 1,200,000 1,000,000 800,000 600,000 400,000 200,000 0 1980

1985

1990

1995 2000 Year

2005

2010

2015

Fig. 4. Area of private land under contract in the Conservation Reserve Program (Farm Bill) in Nebraska, USA during 1986–2010. The initial year for the program was 1986 (data from the National Agricultural Statistics Service, USDA). Fig. 4. Área de terrenos privados bajo contrato del Programa de Conservación de Reservas (Conservation Reserve Program, Farm Bill) en Nebraska, USA, durante el periodo 1986–2010. El año de inicio del programa fue 1986 (datos del Servicio Estadístico Nacional de Agricultura, USDA).

neighboring ranches in the Great Plains of the US may each support populations of greater prairie– chickens (Tympanuchus cupido) on grazing lands for cattle (fig. 5A). If the ranchers are individually approached by a company offering to lease lands for wind energy platforms, each rancher might want to maximize the number of turbines on their property because of direct competition with neighbors for a limited number of leases. Each ranch, then, would potentially be host to wind power (fig. 5B), and planning for siting would be conducted on a ranch–by–ranch basis. Some evidence suggests that prairie–chickens avoid large structures on the landscape (Hagen et al., 2011; Pruett et al., 2009), so it is possible that wind development on the four ranches could cause a decline in space available for prairie–chickens (fig. 5B). In addition, the access roads required for the wind development could also reduce the grazing capacity on each ranch. An alternative scenario would be for the four ranches to form a common–interest community with the purpose to provide more effective planning and profit–base from wind energy, wildlife–based enterprises, and cattle. The results of a joint effort to find the most appropriate location for wind energy could allow the concentration of wind platforms on one section of the association’s lands, which would leave the majority of the prairie–chickens on the lands unaffected by foreign structures. The ranch might be able to develop a rotational grazing schedule that could allow them to maintain stocking levels, across all ranches, close to the pre–association levels (fig. 5C).

Benefits of scale to investors Private reserves will survive as long as private landowners can maintain economic benefits. As investments, conservation done in this manner has the potential to pay for itself, but this demands that landowners have the training and education needed to make good decisions. Marketing strategies for ecotourism can be conducted more effectively and efficiently on behalf of a set of landowners with a large land base than for a single, smaller property (Powell, 2010). If separate landowners are competing for limited tourists, each must produce marketing materials, maintain web sites, attend expositions, and provide staff to make reservations. An association of landowners can reduce these costs by cooperating. With a more diverse landscape (a better product) to market, an association may also realize more income (Naidoo et al., 2011). Last, landowners may also find NGOs and management agencies willing to provide more time and expertise to facilitate management plans, given the history of decisions of the landowner group (Powell, 2010). The association offers the advantage of a single contact point, and a mechanism to develop one management plan that impacts multiple farms or ranches. Case studies Conservation through common interest communities on private reserves is a model that should be consi-

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Fig. 5. Depictions of the distribution of potential sources of revenue on four ranches in the Nebraska Sandhills region: A. Status quo, with cattle grazing and greater prairie–chicken populations on each of the ranches; B. Introduction of wind energy development on the four competing ranches; livestock stocking is potentially reduced, and prairie–chickens could be relegated to areas away from turbines (see text); C. Distribution of elements in B, but in the context of a common–interest community that optimizes wind energy development and cattle grazing, which allows for maintenance of prairie–chicken populations. Dotted lines show property boundaries, but allow flow of income and expenses among ranches. Fig. 5. Descripciones de la distribución de las fuentes potenciales de ingresos de cuatro ranchos en la región de Nabraska Sandhills: A. Status quo, con ganado pastoreando y poblaciones de gallos de las praderas grandes en cada uno de los ranchos; B. Introducción de instalaciones de energía eólica en los cuatro ranchos; se observa una reducción potencial del ganado, y los gallos de las praderas podrían quedar relegados a zonas lejanas a las turbinas (véase el texto); C. Distribución de los elementos en B, pero en el contexto de una comunidad de intereses comunes que optimice el desarrollo de la energía eólica y el pastoreo del ganado, lo que permite el mantenimiento de las poblaciones de gallos de las praderas. Las líneas de puntos son los límites de las propiedades, pero permiten el flujo de entradas y salidas entre los ranchos.

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dered, especially when ecotourism efforts can result in meeting biodiversity or population goals for species of interest (Naidoo et al., 2011). Ecotourism is built on the notion that value can be realized from wildlife and landscapes (Freese, 1998). The following case studies support the theory that conservation can be achieved through the motivations of individual landowners, when the appropriate structure is in place to empower them. Other examples of common interest communities exist throughout the world, especially in Australia, western North America, and central and southern Africa (Schutz, 2010); these two case studies provide details for contrasting examples on two continents. Freehold conservancies in Namibia An example of a landscape–scale management system can be found in the grasslands and shrublands of Namibia, in southern Africa, where cattle farmers have joined together to form conservancies. Before conservancies were established, many farmers built 2–m game fences to restrict the flow of large, game animals. Conservancies provided a mechanism for neighbors to benefit from an integrated landscape (Shaw & Marker, 2011). Namibian landowners form agreements with neighbors about consumptive use limits, habitat management, water management, and ecotourism development. Namibian conservancies have from 5 to 58 farms and range from 75,650 to 500,000 ha; size is generally limited, socially, by distances that neighbors are comfortable driving for meetings (Powell, 2010). Namibia is now home to 23 private conservancies, which are registered with the Ministry of Environment and Tourism. Each conservancy must have a constitution, which defines the relationship among its members and outlines its initial management plan. Conservancies may negotiate with the Ministry to become exempt from typical game permits and use restrictions (Shaw & Marker, 2011). Most conservancies charge member fees to support basic operation or conservation efforts, either on a per hectare or per member basis (Powell, 2010). Namibia’s conservancies each have a distinct flavor because of the heritage of their members and the landscapes in which they exist. Wildlife conservation and poaching protection are primary goals, which contribute to conservation efforts. But, members also list social networking as a goal, which indicates the importance of communication and trust between members. Last, and perhaps realistically, a goal of conservancies is profit. Powell (2010) quoted one conservancy officer, reflecting on their membership: ‘In their eyes, the conservancy will only be valuable for them if the conservancy can increase their profit.’ Namibia’s conservancies also exist across a gradient from arid to semi–arid to more mesic conditions. As the land becomes more productive (better soils, more precipitation), tradeoffs occur in profitability of potential ventures (fig. 5C; Brown, pers. comm., Namibian Nature Foundation). Wil-

Powell

dlife in Namibia are uniquely adapted to more arid zones (relative to domestic animals), and tend to be preferred as an investment in that environment; some landowners in arid regions of Namibia have removed all livestock from their farms in favor of 'farming with wildlife'. However, in more productive zones, cattle co–exist with wildlife, because of the economic return that is available from livestock (fig. 2; Powell, 2010). Such a gradient creates contrasting landscapes in which for biologists to engage landowners; namely, private reserves and other conservation efforts may be easier to develop in regions with less productive lands (fig. 2). Row crops are not common in Namibia, but biologists in regions that can support row crops will encounter a situation in which the conservation trade–offs are further complicated by the high potential for return from production agriculture. Biologists are very aware of the geographic location of thresholds at which grazing becomes feasible (fig. 2A) and at which row–crop agriculture becomes more profitable than grazing (fig. 2B). Greater Gracie Creek Landscape An example of the emerging nature–based entrepreneurship on private reserves can be found near on a 4,800–ha ranch near Burwell, Nebraska, USA. In 2001, the younger generation of the Switzer family voiced an interest to return to the family’s cattle ranch, yet economic reality demonstrated that such a decision was impossible without additional ventures. The family began to diversify their cattle ranch by building a lodge and offering bird watching, boating, guided hunting, and horseback riding. The family found economic value in the leks (breeding grounds) of sharp–tailed grouse (Tympanuchus phasianellus) and greater prairie–chickens, which they now share with their visitors during March and April each spring. The family business, Calamus Outfitters, provided initial opportunities for the second generation to live on the ranch, but the venture’s success was limited by the size of the ranch (Sortum, 2011). Recently, the Switzers joined with two neighboring ranches to form general agreements regarding access and use. The three ranches, as newly branded Greater Gracie Creek Landscape, have become the first private land area in Nebraska to be designated an Important Bird Area by the Audubon Society. The joint group also allows the Switzer’s to market their neighbors’ special beef, known as Morgan Ranch American Wagyu Kobe (Sortum, 2011). To date, the agreement between the Switzers and their neighbors has not officially reached the level of a legal association described as the common–interest community (Schutz, 2010), but those discussions continue. The Switzers have become known as advocates for grassland conservation in the region, and will soon host the first annual Prairie Chicken Festival to showcase educational and recreational activities on their ranch. Regardless of their fondness for conservation, the reality of private lands conservation is expressed in their statement: 'If it pays, it stays'.

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Conclusion Common–interest communities, such as Namibia’s conservancies and the fledgling associated ranches in Nebraska, can provide incentive and scale for effective wildlife management. Private lands biologists should consider the potential for private investment to fuel conservation efforts that can be long–lasting and robust to changing economic and political environments. Management agencies and NGOs must train biologists to facilitate multi–owner groups to promote cross– property agreements for private reserves. The legal means to such ends will vary around the globe; in the US, the simple agreement used to form common–interest communities such as housing and lake associations can be applied in rural settings (Schutz, 2010). The toolbox available to private lands biologists will continue to include, in some form, agriculture subsidies, tax incentives, and conservation easements. But, it is time to embrace opportunities that exist on private reserves. Coordinating and facilitating the development of private reserves in the context of a common–interest community is not easy, as it involves managing people (Powell, 2010). But, Knight (1999) argued that the easy steps in conservation have been taken, and the future will involve many tough conversations and investments of time and energy to make advances in conservation on private properties. Acknowledgements I am grateful for financial support from the University of Nebraska–Lincoln’s (UNL) Research Council, the Grassland Foundation, and World Wildlife Fund. I was supported by a Fulbright Fellowship through the US State Department and the Polytechnic of Namibia (Department of Nature Conservation) while these concepts were developed. C. Brown, Namibian Nature Foundation, provided initial depiction of portions of Figure 2. D. Ebbeka, UNL School of Natural Resources, created Figure 5. An earlier version of this manuscript was improved by reviews by M. Bomberger Brown, B. Franzetti, two anonymous reviewers, and the Associate Editor. This research was supported by Hatch Act funds through the University of Nebraska Agricultural Research Division, Lincoln, Nebraska. References Barbarika, A., 2009. Annual summary and enrollment statistics: FY2009. Natural Resources Analysis Group, Economic and Policy Analysis Staff, Farm Service Agency, U.S. Department of Agriculture: Washington, DC. <http://www.apfo.usda.gov/ Internet/FSA_File/fyannual2009.pdf> Accessed 2 February 2012. Barten, P. K., Damery, D., Catanzaro, P., Fish, J., Campbell, S., Fabos, A. & Fish, L., 2001. Massachusetts family forests: birth of a landowner cooperative. Journal of Forestry, 99: 23–30. Brown, R., 2010. A conservation timeline: milestones

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Perdix XIII Comitè científic / Comité científico / Scientific Comittee Species extinctions and population dynamics Christos Thomaides Technological Education Inst. of Lamia, Karpenisi, Greece Wildlife law and policy: Markus Jenny Schweizerische Vogelwarte, Switzerland Conservation and management of migratory gamme species Christos Thomaides Technological Education Inst. of Lamia, Karpenisi, Greece Wildlife biology, behaviour and game species management Francis Buner The Game and Wildlife Conservation Trust, UK Julie Ewald The Game and Wildlife Conservation Trust, UK Sandor Faragó Inst. of Wildlife Management and Vertebrate Zoology, Univ. of West Hungary, Hungary Elisabeth Bro Office National de la Chasse et de la Faune Sauvage, ONCFS, France Manel Puigcerver Univ. de Barcelona, Spain José Domingo Rodríguez–Teijeiro Univ. de Barcelona, Spain Christos Thomaides Technological Education Inst. of Lamia, Karpenisi, Greece Methodologies, models and techniques: Carlos Sánchez Univ. de León, Spain Human dimensions of game wildlife management: Julie A. Ewald The Game and Wildlife Conservation Trust, UK

Assessors dels articles / Asesores de los artículos / Referees of papers Nicholas J. Aebischer The Game and Wildlife Conservation Trust, UK Mª Victoria Arruga Univ. de Zaragoza, Spain Mathieu Boos Research Agency in Applied Ecology, France Elisabeth Bro Office National de la Chasse et de la Faune Sauvage, France Stephen Browne Fauna & Flora International (Singapore) Ltd, Singapore John P. Carroll Warnell School of Forestry and Natural Resources, Univ. of Georgia, USA Ralph Dimmick Univ. of Tennessee, USA Roger Draycott The Game and Wildlife Conservation Trust, UK Jerome Duplain Schweizerische Vogelwarte, Switzerland Sandor Faragó Inst. of Wildlife Management and Vertebrate Zoology, Univ. of West Hungary, Hungary Jean–Carles Guyomarc'h Univ de Rennes I, France Eleftherios Hadjisterkotis The Ministry of the Interior, Nicosia, Cyprus Markus Jenny Schweizerische Vogelwarte, Switzerland Alberto Meriggi Univ. di Pavia, Italy Florian Millot Office National de la Chasse et de la Faune Sauvage, ONCFS, France Marek Panek Polish Hunting Association, Research Station, Poland José Antonio Pérez Garrido Univ. de León, Spain François Ponce–Boutin Office National de la Chasse et de la Faune Sauvage, ONCFS, France Dick Potts England, UK Manel Puigcerver Univ. de Barcelona, Spain Ettore Randi Inst. Nazionale per la Fauna Selvatica (INFS), ISPRA, Italy François Reitz Office National de la Chasse et de la Faune Sauvage, ONCFS, France José Domingo Rodríguez–Teijeiro Univ. de Barcelona, Spain Carlos Sánchez, Grupo de Producción y Gestión Cinegética, Univ. de León, Spain Jörg E. Tillmann Inst. für Wildtierforschung an der Stiftung Tierärztliche Hochschule, Germany Philip Warren The Game and Wildlife Conservation Trust, UK

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Perdix XIII Índex / Índice / Contents Species extinctions and population dynamics 311–319 S. Faragó, G. Dittrich, K. Horváth–Hangya & D. Winkler Twenty years of the grey par tridge population in the LAJTA Project (Western Hungary) 321–331 P. J. K. McGowan, L. L. Owens & M. J. Grainger Galliformes science and species extinctions: what we kwow and what we need to know

Conservation and management of migratory game species 333–342 J. D. Rodríguez–Teijeiro, F. Sardà–Palomera & M. Puigcerver Post–breeding movements and migration patterns of western populations of common quail (Coturnix coturnix): from knowledge to hunting management 343–352 M. Puigcerver, F. Sardà–Palomera & J. D. Rodríguez– Teijeiro Deter mining population tr ends and conservation status of the common quail (Coturnix coturnix) in Western Europe

Wildlife biology, behaviour and game species management 353–362 N. J. Aebischer & J. A. Ewald The grey partridge in the UK: population status, research, policy and prospects 363–369 J. A. Ewald, G. R. Potts & N. J. Aebischer Restoration of a wild grey partridge shoot: a major development in the Sussex study, UK 371–380 V. A. Bontzorlos, C. G. Vlachos, D. E. Bakaloudis, E. N. Chatzinikos, E. A. Dedousopoulou, D. K. Kiousis & C. Thomaides Rock partridge (Alectoris graeca graeca) population density and trends in central Greece

381–386 R. A. H. Draycott Restoration of a sustainable wild grey partridge shoot in eastern England 387–393 K. Buckley, P. Kelly, B. Kavanagh, E. C. O’Gorman, T. Carnus & B. J. McMahon Every partridge counts, successful techniques used in the captive conservation breeding programme for wild grey partridge in Ireland 395–404 A. Mateo–Moriones, R. Villafuerte & P. Ferreras Does fox control improve red–legged partridge (Alectoris rufa) survival? An experimental study in Northern Spain 405–413 E. Bro, P. Mayot & F. Reitz Effectiveness of habitat management for improving grey partridge populations: a BACI experimental assessment 415–418 D. A. Butler, W. E. Palmer & M. P. Cook The invertebrate diet of northern bobwhite chicks in Georgia, United States

Methodologies, techniques

models

and

419–428 T. Liukkonen, L. Kvist & S. Mykrä Microsatellite markers show distinctiveness of released and wild grey partridges in Finland 429–435 P. Tizzani, E. Negri, F. Silvano, G. Malacarne & P. G. Meneguz Does the use of playback affect the estimated numbers of red–legged partridge Alectoris rufa

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Twenty years of the grey partridge population in the LAJTA Project (Western Hungary) S. Faragó, G. Dittrich, K. Horváth–Hangya & D. Winkler

Faragó, S., Dittrich, G., Horváth–Hangya, K. & Winkler, D., 2012. Twenty years of the grey partridge population in the LAJTA Project (Western Hungary). Animal Biodiversity and Conservation, 35.2: 311–319. Abstract Twenty years of the grey partridge population in the LAJTA Project (Western Hungary).— The Lajta Project covers 3,065 ha. Within this area crop cultivation is dominant. Fields are separated from each other by forest belts and tree rows, extending altogether over roughly 120 ha. This habitat structure characterized by cultivation of 12–15 field crops sustained partridge population with densities of 1.75 birds/km2 (1991). The Project started in 1991/1992 and aimed to increase the carrying capacity for grey partridge and other small game species living in the area. A full–time gamekeeper was employed and habitat improvements were initiated. Four years later, the breeding population increased to 10.1 birds/km2. Besides increased numbers of nesting pairs, the number of reared chicks also increased, from 5.1–11.2 individuals/km2 in 1990 to 27.3–38.4 individuals/km2 in 1994. However, field sizes did not change significantly. Although the lengths of field margins increased by approximately 25% (from 82 m/ha to 115 m/ha) under the influence of habitat management, they still reached only half those found in the countries of Central Europe where private ownership of land properties is dominant. After the privatisation of fields in 1995 as part of the political change in Hungary —affecting approximately 50% of the project area— the possibilities of habitat improvement decreased, and the technological pressure on large–scale farming area increased. Following these processes the grey partridge population again decreased to 1.43 birds/ km2 in 1997. As a result of the new management strategy applied in the project since 1996 we observed a slow increase in the breeding population, which stabilized at around 5 birds/km2, between 2007 and 2009. The August density increased in the same period from 4.5 birds/km2 to 13–17 birds/km2. During the two decades in which this research was conducted, chick mortality and winter mortality were extremely high. The key factors influencing grey partridge population dynamics in our study area seem to be clutch and chick losses and winter mortality. To determine the relationship between environmental factors and the grey partridge population parameters, principal component analysis (PCA) was used. August grey partridge density was positively associated with ecotone density and bags of red fox and avian predators (magpie and hooded crow). The density of grey partridge in spring and August also correlated with feral dog bag and feral cat bag. Habitat parameters showed a positive correlation with the density of grey partridge both in spring and in August. The levels of abundance of feral dog and feral cat populations also correlated negatively with winter losses. Key words: Grey partridge, Long–term monitoring, Predators, Habitat, LAJTA Project, Hungary. Resumen Veinte años de la población de perdiz pardilla en el Proyecto LAJTA (Hungría occidental).— El Proyecto LAJTA cubre 3.065 ha. En esta área dominan las tierras de cultivo. Los campos están separados entre sí por cinturones forestales e hileras de árboles, que tienen en conjunto una superficie aproximada de 120 ha. Esta estructura de hábitat, caracterizada por el cultivo de 12–15 cosechas, mantenía una población de perdices con densidades de 1,75 aves/km2 (1991). El Proyecto empezó en 1991/1992 con el propósito de aumentar la capacidad de carga para la perdiz pardilla y otras especies de caza similares que vivían en la zona. Se contrató a un guardabosques a jornada completa y se inició la mejora del hábitat. Cuatro años más tarde, la población de cría había aumentado a 10,1 aves/km2. Además del aumento de parejas nidificantes, también aumentó la cantidad de pollos criados, de 5,1–11,2 individuos/km2 en 1990 a 27,3–38,4 individuos/km2 en 1994. No obstante, el tamaño de los campos no cambió significativamente. Aunque la longitud de los márgenes de los campos aumentó aproximadamente en ISSN: 1578–665X

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un 25% (de 82 m/ha a 115 m/ha) por la influencia de la gestión del hábitat, solamente se alcanzó la mitad de los que se encuentran en los países de Europa central, donde predomina la propiedad privada de la tierra. Tras la privatización de los campos en 1995, como parte del cambio político húngaro —lo que afectó aproximadamente al 50% del área del proyecto— las posibilidades de mejorar el hábitat disminuyeron, y aumentó la presión tecnológica sobre la agricultura a gran escala. Como consecuencia de este proceso la población de perdiz pardilla volvió a decrecer hasta 1,43 aves/km2 en 1997. Como resultado de una nueva estrategia de gestión aplicada al Proyecto desde 1996, observamos un lento incremento de la población de cría, que se estabilizó alrededor de 5 aves/km2 del 2007 al 2009. La densidad en el mes de agosto aumentó durante el mismo periodo de 4,5 aves/km2 a 13–17 aves/km2. Durante las dos décadas en las que se llevó a cabo este estudio, la mortalidad de las crías y la mortalidad invernal eran extremadamente altas. Los factores clave que influencian la dinámica de poblaciones de la perdiz pardilla en nuestra área de estudio parecen ser las pérdidas de huevos y pollos, y la mortalidad invernal. Con el fin de determinar la relación entre los factores ambientales y los parámetros de la población de la perdiz pardilla, se utilizó un Análisis de Componentes Principales (ACP). La densidad de la perdiz pardilla en el mes de agosto se asociaba positivamente con la densidad del ecotono y la caza de zorros comunes y otros depredadores de aves (cornejas cenicientas y urracas). La densidad de las perdices también se correlacionaba en primavera y agosto con la caza de gatos y perros cimarrones. Los parámetros del hábitat tenían una correlación positiva con la densidad de la perdiz pardilla tanto en primavera como en agosto. Los niveles de abundancia de las poblaciones de gatos y perros cimarrones también se correlacionaron negativamente con las pérdidas invernales. Palabras clave: Perdiz pardilla, Monitorización a largo plazo, Predadores, Hábitat, Proyecto LAJTA, Hungría. Received: 22 XII 11; Conditional acceptance: 10 II 12; Final acceptance: 30 IV 12 Sándor Faragó, Gábor Dittrich, Katalin Horváth–Hangya & Dániel Winkler, Inst. of Wildlife Management and Vertebrate Zoology, Univ. of West Hungary, Fac. of Forestry, H–9400 Sopron, Ady Endre út 5, Hungary. Corresponding author: S. Faragó. E–mail: farago@emk.nyme.hu

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Introduction

such as natality, chick survival rate and mortality changed? What are the mortality periods that basically determine population dynamics? What environmental parameters influence the population change, and what are the most important ones?

The decrease of farmland wildlife and the negative impact of intensive agricultural technologies has become evident over the last decades (Donald et al., 2001; De Leo et al., 2004; Verhulst et al., 2004; Vickery et al., 2004.) The change has particularly influenced grey partridges (Perdix perdix), which are the most sensitive species in the areas and their populations have decreased throughout Europe (Bro et al., 2000; BirdLife International, 2004; Kuijper et al., 2009). The situation is no different in Hungary (Faragó, 1988; Báldi & Faragó, 2007). When studying this phenomenon it is clear that the extent of the changes differs in the various countries of Europe depending on the ecological conditions, political factors and economic possibilities. As a result, a research and management program was started in Hungary to maintain farming and support the conservation of grey partridge, in particular (Faragó, 1997a, 1998). The LAJTA Project became the flagship of the Hungarian Partridge Conservation Program. We started the LAJTA Project in 1992, in order to implement complex monitoring of sma ll game species and their environment (Faragó, 1991). The key species of this project is the grey partridge because it is the indicator species of the farmlands and the changes in its population reflect the positive and negative effects in the environment in the fastest way (Faragó, 1997b; Faragó & Buday, 1998). In the Project’s operation we followed the newly composed (Robertson, 1991) directives of wise use involving habitat development and predator control. We wanted to answer the following questions: how has the size/density of the grey partridge population changed? How have parameters of the population

Methods Study area The area of the LAJTA Project is 3,065 ha in Kisalföld (Little Hungarian Plain), Western Hungary (fig. 1). Until 1995, the Lajta–Hanság Co. had managed the area exclusively. However, in 1995, due to compensations/ privatization, 50% of the area was in the hands of smallholders. This area has a continental climate (mean annual temperature is 9.6oC, annual precipitation is 504 mm, mean relative humidity is 73%) where the main crops are grain and maize. About 40% of the farming is large scale (Lajta–Hanság Joint Stock Company, average field size 50 ha) and 60% is small scale (small holders, average field size 2 ha). In both cases, there is intensive technology, which, from the point of view of mechanization and the use of chemicals, has not changed in the past twenty years. Fields are separated from each other by forest belts, tree rows and similar, extending over roughly 120 ha. Pasturing did not take place in the Project territory and the fodder demand of animal husbandry was supplied by growing alfalfa and silo maize. Study design During the study we used the following methods: for the grey partridge population we used total population

N Mosonszolnok

Slovakia

Ukraine

Austria

Varvalog

LAJTA Project

Hungary Slovenia Croatia

Fig. 1. Location of the study area in Hungary. Fig. 1. Localización del área de estudio en Hungría.

Serbia and Montenegro

Romania

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1,200

Number

900

600

Spring

2010

2009

2008

2007

2006

2005

2004

2003

2002

2001

2000

1999

1998

1997

1996

1995

1994

1993

1992

1991

300

August

Fig. 2. Dynamics of the grey partridge population in the LAJTA Project. Fig. 2. Dinámica de la población de perdiz pardilla en el Proyecto LAJTA.

assessment with a mapping method during weekly records. In the case of birds of prey (marsh harrier, Circus aeruginosus; goshawk, Accipiter gentilis), red fox (Vulpes vulpes) and European badger (Meles meles) we mapped nests and burrows. We carried out total population assessments every two weeks to calculate the number of birds of prey. We also calculated the monthly average. We recorded/mapped habitats on the 15th day of each month, and we also mapped habitat–improvements. Every year, the following habitat improvements are carried out in the LAJTA Project: chemical–free field margins, weedy strips between two crop fields, harrowed strips and unmown margins of grassland and alfalfa fields were left; mowing of cereal field margins was postponed. The partridge fields were set aside (Faragó & Buday, 1998). The game–keeper of the Project continuously controls predators, especially before and during the breeding period of grey partridge. We collected climatic data from the meteorological station of Mosonmagyaróvár City. The game–keeper of the Project keeps a record of the predator control data. Data analyses The chick survival rate was calculated using Potts’ model (1986). We determined the most important mortality period by key–factor analysis (Chlewski &

Panek, 1988). Farmland diversity was calculated by applying the Shannon–Wiener diversity index H’ to measure the composition of the landscape, using the territorial data of the land cover types. To reduce the number of variables in the environmental data matrix, principal component analysis (PCA) was performed on 14 selected variables including habitat, predator and meteorological variables and excluding variables which had no equal data (fox burrows and birds of pray monitoring results). All variables were tested for normality and homogeneity of variances. Only PCA factors with eigenvalues more than 1.0 were selected (Kaiser Criterion). Factor loadings were rotated with a varimax raw transformation. Next, linear regression was performed on the principal components derived to examine the correlation between the components describing the environment and grey partridge population data (spring and August density, chick survival rate and key factors). All statistical analyses were carried out using the SPSS (SPSS, 1999). Results The grey partridge population The original habitat structure was able to support the grey partridge population with densities of 1.8 birds/km2

Animal Biodiversity and Conservation 35.2 (2012)

The most important ecological parameters Although the above statements define the strength and timing of mortality factors, they do not focus on their cause. Only detailed ecological investigations can bring this to light. The mid–month habitat mapping clearly shows the dynamic change from month to month in the habitat structure due to the agricultural growing cycles and technological processes. The changes in the May positions recorded during the 20 years showed that the dominance of spring crops in the early 2000s was replaced by the dominance of autumn crops with the low and nearly permanent level of perennial plants. From the point of view of nesting possibilities for grey partridge, we may evaluate this kind of process as advantageous. Habitat diversity has increased in the last five years, after a slow 15–year decrease (fig. 4). This can be explained by the increased number of cultivated plants and by the relations of an equal dominance. Another important characteristic of the habitat quality is the length and the density of ecotones. Due to the decrease in field size, the density of ecotones grew 65% from 82 m/ha to 135 m/ha. The habitat improvement resulted in a further increase with annual values of 11–45 m/ha. Consequently, the ecotone density increased from 106 m/ha to 174 m/ha by 64% during the 20 years (fig. 5). Predator pressure is one of the most important factors influencing the grey partridge population. The number of nesting pairs of raptors in the Project showed a decreasing or stagnant tendency, and the same occurred in the case of hooded crow (Corvus cornix) –13→1–2 pairs —decreasing, and magpie (Pica pica), 1–3 pairs— stagnant. Based on the breeding and non–breeding raptor censuses carried out every two weeks in the

0.8

r = 0.726

0.6 0.4 0.2 0

0.5

0.8

1.5

r = 0.631

0.6 0.4 0.2 0 0 0.8 0.6 k3

(1991). After four years, the breeding population increased to 10.1 birds/km2. Besides increased numbers of nesting pairs, the number of reared chicks increased from 5.1–11.2 individuals/km2 in 1990, to 27.3–38.4 individuals/km2 in 1994. After field privatization in 1995 as part of the political changes in Hungary affecting approximately 50% of the project area, the possibilities of habitat improvement decreased, and the agricultural technological pressure of large scale farming area increased. At the same time, the grey partridge population decreased again to 1.4 birds/km2 in 1997. With the management strategy applied in the project since 1996, the breeding population increased slowly and stabilized at around 5 birds/km2 between 2007 and 2009. The August population density increased in the same period from 4.5 to 13–17 birds/km2 (fig. 2). During the last two decades, egg, chick and winter losses were extremely high. The values of the former changed between 44–90%; winter population declined between 30–81%. The losses of adult birds in summer were lower, between 7–51%. Chick survival rate changed in inverse ratio with egg and chick losses. According to the key factor analysis, the most important mortality factors influencing the grey partridge population dynamics seem to be clutch and chick losses (k1) and winter mortality (k3) (fig. 3).

315

0.5 r = 0.428

0.4 0.2 0

0.5

Fig. 3. Correlations between three individual mortalities and total mortality. Fig. 3. Correlaciones entre tres mortalidades individuales y la mortalidad total.

LAJTA Project, only the marsh harrier, the hen harrier (Circus cyaneus) and the goshawk can be regarded as effective predators of the grey partridge in Hungary (Faragó, 2002). The goshawk can be observed in 1–2 exemplars all year round (max. 0.06 birds/km2). During the breeding period (April–August), the marsh harrier appears with increasing density (0.16–0.45 bird/km2), although its nesting was not demonstrated in the Project. In the same way, the winter permanent appearance of the hen harrier is important (average of 0.19 birds/km2). Due to the work of the game–keeper, the number of inhabited burrows of the key predator, red fox, decreased from the initial 27 burrows (0.9 burrow/km2) to 5 burrows (0.16 burrow/km2). Nevertheless, the European badger population showed a slow increase despite the fact that in 2001 this formerly protected species returned to huntable status again. The density of inhabited European badger burrows was 0.23 burrows/km2 in 2010, exceeding that of red fox. The regulation of six predator species is important. There was an increase in the bags of the red fox, magpie, hooded crow and Eurasian jay (Garrulus glandarius). The feral dog population decreased while the feral cat

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3,000 2,500

2,000 1,500 1,000 500

Perennial plants

Winter crops

2010

2009

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2006

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2004

2003

2002

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Spring crops

2.50 2.00

1.50 1.00 0.50

2010

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2004

2003

2002

2001

2000

1999

1998

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Fig. 4. Dynamics of yearly habitat availability and habitat diversity in the breeding season in the LAJTA Project. Fig. 4. Dinámica de la disponibilidad anual de hábitat y diversidad de hábitats durante la estación de cría en el Proyecto LAJTA.

population stagnated. Because of the immunization against rabies in 1992, the red fox population has increased everywhere in the Project area. This increment appeared in the population size and later in the bag as well. The immunization may have caused the population rise of protected small mammal predator species, but we do not have any information on this. On the whole, both the censuses and the bag records show predator pressure increased. As well as the above–mentioned findings, important variations in weather conditions had to be considered, especially climate change. During the breeding period temperatures below the average (16.3°C) were recorded only in 1997, and in the other years temperatures were sometimes 2°C higher. As compared to the normal 316 mm precipitation, there was often drought; rainfall was only heavy in 1996 and 2010. The winter season was colder than the 3°C average in three years only and was over 3°C higher in other years. The mean winter precipitation (264 mm) occurred only in the winter of 1999/2000 and 2008/2009. In the other years it was always less.

For the partridge, which is very much adapted to the continental climate, the above changes are favourable. In the juvenile/adult ratio, which shows the success of reproduction, the rise of temperature caused a slight increase, while the change in precipitation did not show any significant impact. Connections between grey partridge population and environmental factors The PCA performed on the 14 selected variables yielded five new variables that together explain 76.08% of the total variance (table 1). The first component accounted for 29.22% of the total variance and it was positively correlated with ecotone density and with the bags of red fox, magpie and hooded crow. The second component accounted for an additional 13.43% of the total variance. This component was positively correlated with the bags of feral dog and feral cat. The third component accounted for nearly the same proportion (12.55%) of the total variance as the previous one. With the exception of the ecotone density, all habitat

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180 160 140

(m/ha)

120 100 80 60 40

Basic density (m/ha)

2010

2009

2008

2007

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2005

2004

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2001

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Increase in density (m/ha)

Fig. 5. Long–term change of length of ecotones in the LAJTA Project. Fig. 5. Cambio a largo plazo de la longitud de los ecotonos en el Proyecto LAJTA.

variables made a prominent contribution. This third component axis was positively correlated with habitat diversity and with the proportion of perennial plants and spring crops, while winter crops had a negative score. The fourth and fifth components each accounted for about 10% of the total variance. The fourth component was mostly determined by the meteorological variables for the period October–March while the fifth component showed a stronger correlation with the meteorological parameters of the April–August period. In both cases, average temperature had positive scores and precipitation had negative scores. For further analysis, the final PCA components were used as independent variables in the regression analyses with grey partridge population parameters as response variables. Several significant models were obtained. August grey partridge density was positively associated with PCA factor 1 (R2 = 0.368; F = 9.901; p < 0.01), a component mainly determined by the ecotone density and bags of red fox and avian predators (magpie and hooded crow). Grey partridge density was also correlated with PCA factor 2 mainly determined by the feral dog bag and the feral cat bag (model for spring density: R2 = 3.481; F = 9.076; p < 0.01; model for August density: R2 = 0.210; F = 4.512; p < 0.05). PCA factor 3, mostly characterized by habitat parameters, showed a positive correlation with both the spring and August grey partridge density (model for spring density: R2 = 0.220; F = 4.808; p < 0.05; model for August density: R2 = 0.208; F = 4.457; p < 0.05).

The effect of feral dog and feral cat populations also manifested in the next model, forasmuch as winter losses (K3) it was negatively correlated with PCA factor 2 (R2 = 0.251; F = 5.357; p < 0.05). Discussion Our investigations suggest a strong link between habitat diversity and population density of grey partridge in spring and autumn. Earlier analysis (Báldi & Faragó, 2007) in regard to Hungarian small game populations characteristically showed the decrease of farmland diversity. Manifold investigations have justified that farmland heterogeneity is a key factor in the maintenance of farmland biodiversity (Benton et al., 2003; Verhulst et al., 2004). In an agricultural environment managed in an intensive way as in the LAJTA Project, there is an opportunity to maintain the grey partridge population if habitats in the territory are improved. If we can increase crop diversification and reduce pesticide use, this will increase the carrying capacity of the territory (Sotherton, 1991; Henderson et al., 2009). It might help if we can also apply the wide spectrum of agri–environment schemes (arable flora management, beetle banks, conservation headlands, crop management, field corner management, grass strips, grassland and scrub management, spring cropping, wild bird cover and overwinter stubble) (Ewald et al., 2010). In general, we can say that a framework of ecologically enhanced

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Table 1. Factor loadings for the first five principal components in PCA on the environmental variables used. Tabla 1. El peso de cada factor (factor loadings) para las cinco primeras componentes principales en el ACP sobre las variables ambientales utilizadas.

Factor 1

Factor 2

Factor 3

Factor 4

Factor 5

Habitat diversity

0.254

0.228

0,806

0.118

0.091

Perennial plants

–0.158

0.099

0,772

–0.073

–0.245

Winter crops

0.367

0.025

–0,755

0.295

–0.119

Spring crops

–0.322

0.172

0,589

–0.351

0.325

Density of ecotones

0.967

-0.031

0,191

–0.022

0.109

Red fox bag

0.842

0.197

–0,117

0.052

0.02

Feral dog bag

–0.225

0.907

0,181

–0.03

–0.026

Feral cat bag

0.273

0.816

0,216

0.242

–0.106

Magpie bag

0.871

-0.093

–0,076

0.239

0.168

Hooded crow bag

0.779

-0.041

–0,101

–0.064

0.015

Average temperature (Apr.–Aug.)

0.379

-0.068

–0,035

–0.007

0.852

Precipitation (Apr.–Aug.)

0.398

0.02

0,243

–0.033

–0.614

Average temperature (Oct.–March)

–0.029

0.195

–0,156

0.873

0.125

Precipitation (Oct.–March)

0.023

0.323

–0,283

–0.606

0.168

Eigen values

4.09

1.88

1.76

1.53

1.39

Explained variance %

29.22

13.43

12.55

10.94

9.94

Cumulated variance %

29.22

42.65

55.20

66.14

76.08

areas is a key habitat structure for grey partridges (Buner et al., 2005). Earlier, the use of a set–aside system was a tool for this implementation. However, in recent years, processes in the agricultural policy of EU have been less favorable. Little emphasis is given to the importance of preserving the carrying capacity and ecotones —such as grassy strips, weedy strips, tree and shrub rows, forest belts and similar. (Faragó, 1998)— and maintaining and protecting wasteland patches (Šálek et al., 2004). Another important question of habitat management is the impact of predators on grey partridge population (Potts, 1986; Kalchreuter, 1991; Tapper et al., 1996; Bauer & Berthold, 1997; Faragó, 1997c; Potts, 2009; etc.). Our results suggest that the key predator red fox has indeed a negative effect on population density. It is also confirmed that the other predators can cause mortality in different periods. Grey partridge population dynamics are generally are determined by many different factors such as economic possibilities, agriculture policies and climatic factors. To adapt to or compensate for these factors, habitat management and predator control should be increased and made more effective in the future in the LAJTA Project.

Acknowledgements We wish to thank LAJTA–HANSÁG Co for all the support they provided throughout the twenty years of the project, and also the Ministry of Rural Development for supporting our work for fifteen years. We also thank TÁMOP 4.2.1.B within the Sopron Regional University Knowledge Centre for the grant that made our research possible in the last two years. References Bauer, H–G. & Berthold, P., 1997. Die Brutvögel Mitteleuropas. Bestand und Gefährdung. Aula–Verlag, Wiesbaden. Báldi, A. & Faragó, S., 2007. Long–term changes of farmland game populations in a post–socialist country (Hungary). Agriculture, Ecosystems & Environment, 118: 307–311. Benton, T. G., Vickery, J. A. & Wilson, J. D., 2003. Farmland biodiversity: is habitat heterogeneity the key? Trends. Ecol. Evol., 18: 182–188. BirdLife International, 2004. Birds in Europe: population estimates, trends and conservation status.

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Cambridge, UK: BirdLife International., BirdLife International Series No. 12. Bro, E., Sarrazin, J. C. & Reitz, F., 2000. Demography and the decline of the grey partridge (Perdix perix) in France. Journal of Applied Ecology, 27: 432–448. Buner, F., Jenny, M., Zbinden, N. & Naef–Daenzer, B., 2005. Ecologically enhanced areas – a key habitat structure for re–introduced grey partridges Perdix perdix. Biological Conservation, 124: 373–381. Chlewski, A. & Panek, M., 1988. Population dynamics of the partridge on hunting grounds of Czempin, Poland. Common Partridge International Symposium, Poland’85: 143–156. De Leo, G. A., Focardi, S., Gatto, M. & Cattadori, I. M., 2004. The decline of the grey partridge in Europe: comparing demographies in traditional and modern agricultural landscapes. Ecological Modelling, 177: 313–335. Donald, P. F., Green, R. E. & Heath, M. F., 2001. Agricultural intensification and the collapse of Europe’s farmland bird population. Proc. R. Soc. London B, 268: 25–29. Ewald, J. A., Aebischer, N. J., Richardson, S. M., Grice, P. V. & Cooke, A. I., 2010. The effect of agri–envirinment schemes on grey partridges at the farm level in England. Agriculture, Ecosystems & Environment, 138: 55–63. Faragó, S., 1988. Die Gestaltung der Bestände des Rebhuhnes und die Lage dieser Vogelart in Ungarn im Jahre 1985. Common Partridge International Symposium, Poland'85: 185–198. – 1991. Über eine Untersuchung im Nachbargebiet – Das LAJTA–Project. BFB–Bericht, 77: 77–84. – 1997a. The Hungarian Partridge Conservation Program. Management and research. Hungarian Small Game Bulletin, 1: 19–30. – 1997b. Dynamics of the partridge population covered by the LAJTA Project (Western Hungary) 1989–1995. Hungarian Small Game Bulletin, 1: 107–132. – 1997c. Habitat improvement in the small game management. Environmental basis of the sustainable small game management. Mezőgazda Kiadó, Budapest. – 1998. Habitat improvement of Hungarian Partridge populations (Perdix perdix): the Hungarian partridge conservation program (HPCP). Game & Wildlife Sciences, 15: 145–156. – 2002. Hunting zoology. Mezőgazda Kiadó, Budapest. [In Hungarian.] Faragó, S. & Buday, P., 1998. Examinations on grey partridge (Perdix perdix) population and its environ-

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ment covered by the LAJTA Project, 1989–1997. Hungarian Small Game Bulletin, 2: 1–250. Henderson I. G., Ravenscroft, N., Smith, G. & Holloway, S., 2009. Effects of crop diversification and low pesticide input on bird population on arable land. Agriculture, Ecosystems & Environment, 129: 149–156. Kalchreuter, H., 1991. Rebhuhn Aktuell. Verlag Dieter Hoffmann, Mainz. Kuijper, D. P. J., Oosterveld, E. & Wymenga, E., 2009. Decline and potential recovery of the European grey partridge (Perdix perdix) population – a review. European Journal of Wildlife Research, 55: 455–463. Potts, G. R., 1986. The Partridge. Pesticides, predation and conservation. Collins, London. – 2009. Restoring a gray partridge (Perdix perdix) population and the future of predation control. In: Gamebird 2006: Quail VI and Perdix XII: 24–25 (S. B. Cederbaum, B. C. Faircloth, T. M. Terhune, J. J. Thompson & J. P. Carroll, Eds.). Warnell School of Forestry and Natural Resources, Athens, USA. Robertson, P., 1991. Wise use and conservation. Gibier Faune Sauvage, 8: 379–388. Šálek, M., Marhoul, P., Pintíř, J., Kopecký, T. & Slabý, L., 2004. Importance of unmanaged wasteland patches for the grey partridge Perdix perdix in suburban habitats. Acta Oecologica, 25: 23–33. Sotherton, N. W., 1991. Conservation headlands: a practical combination of intensive cereal farming and conservation. In: The Ecology of temperate Cereal Fields: 373–397 (L. G. Firbank, N. Carter, J. F. Derbyshire & G. R. Potts, Eds.). Blackwell Scientific Publications, Oxford. SPSS, 1999. SPSS Base 10.0. SPSS Incorporation, Chicago. Tapper, S. C., Potts, G. R. & Brockless, M. H., 1996. The effect of an experimental reduction in predation pressure on the breeding success and population density of grey partridges Perdix perdix. Journal of Applied Ecology, 33: 965–978. Verhulst, J., Báldi, A. & Kleijn, D., 2004. Relation between land–use intensity and species–richness and abundance of birds in Hungary. Agriculture, Ecosystems & Environment, 104: 465–473. Vickery, J. A., Bradbury, R. B., Henderson, I. G., Eaton, M. A. & Grice, P. V., 2004. The role of agri–environment schemes and farm management practices in reversing the decline of farmland birds in England. Biological Conservation, 119: 19–39.

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Galliformes science and species extinctions: what we know and what we need to know P. J. K. McGowan, L. L. Owens & M. J. Grainger

McGowan, P. J. K., Owens, L. L. & Grainger, M. J., 2012. Galliformes science and species extinctions: what we know and what we need to know. Animal Biodiversity and Conservation, 35.2: 321–331. Abstract Galliformes science and species extinctions: what we know and what we need to know.— In early 2010, the 193 Parties that had signed up to the Convention on Biological Diversity all acknowledged that they had failed to meet the target that they had set themselves in 1992 of significantly reducing species extinctions by 2010. At the end of the year they set a new and more ambitious target of preventing species extinctions by 2020. Achieving that target will require much greater efficiency in the use of resources and research has a very significant role to play in making this happen. There are 290 species of Galliformes of which 26% are considered at risk of extinction, compared with 12% of all 10,000 bird species. At the same time there is significant research literature on the group that stretches back decades for some species. It is timely, therefore, to consider whether it is possible to increase the efficiency and global impact of gamebird research so that, with careful planning that involves more strategic direction and sharing of lessons learnt, game biologists can play a significant role in achieving the 2020 target for species adopted by the Convention on Biological Diversity. Specific areas in need of this lesson sharing approach are population estimation and threat assessment, analysis of exploitation and determining the ecological basis of successful interventions. Key words: Galliformes, Conservation, Policy, Convention on Biological Diversity, Extinction risk. Resumen Extinciones de especies de Galliformes y conocimientos científicos: lo que sabemos y lo que necesitamos saber.— A principios de 2010, las 193 partes que habían firmado el Convenio sobre la Diversidad Biológica reconocieron que no habían cumplido el objetivo que ellas mismas habían fijado en 1992 de reducir de forma significativa las extinciones de especies en 2010. Al final del año establecieron un objetivo nuevo y más ambicioso que consistía en evitar las extinciones de especies en 2020. Lograr dicho objetivo requerirá una utilización mucho más eficiente de los recursos y la investigación tiene un papel fundamental en hacer que esto ocurra. Existen 290 especies de Galliformes, de las cuales el 26% se considera en peligro de extinción, en comparación con el 12% del total de las 10.000 especies de aves. Al mismo tiempo, hay numerosos estudios publicados sobre el grupo que abarcan décadas para algunas especies. Por consiguiente, es oportuno analizar si es posible aumentar la eficiencia y las repercusiones a escala mundial de la investigación sobre aves de caza, de forma que, con la planificación meticulosa que conlleva más orientación estratégica e intercambio de experiencias, los biólogos especializados en este tipo de aves puedan desempeñar una función destacada en la consecución del objetivo de 2020 para las especies aprobado por el Convenio sobre la Diversidad Biológica. Los ámbitos específicos que necesitan este planteamiento de intercambio de experiencias son la estimación de la población y la evaluación de las amenazas, el análisis de la explotación y la determinación de la base ecológica de las intervenciones que hayan obtenido buenos resultados. Palabras clave: Galliformes, Conservación, Políticas, Convenio sobre la Diversidad Biológica, Riesgo de extinción. Received 19 III 12; Conditional acceptance: 25 V 12; Final acceptance: 8 VI 12 Philip J. K. McGowan, Laura L. Owens & Matthew J. Grainger, World Pheasant Association, Newcastle Univ. Biology Field Station, Close House Estate, Heddon on the Wall, Newcastle upon Tyne NE15 0HT, UK. Corresponding author: Philip McGowan. E–mail: director@pheasant.org.uk ISSN: 1578–665X

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Introduction The deteriorating conservation status of the world’s species is well documented (Barnosky et al., 2011). Conservation science is responding to this decline by offering increasing knowledge about species distributions and life history and the pressures on them, the actions that can address these pressures, and how conservation can then be implemented. The global policy context is arguably more sympathetic to species conservation than it has ever been. At the same time, however, public funds that are targeted explicitly for species conservation are being cut on a significant scale. Taken together this means that if those concerned with conservation are to find ways of capitalising on new opportunities to prevent species extinctions, it is imperative to significantly increase the efficiency with which resources are deployed for research, management and monitoring. A fundamental requirement for greater efficiency is increased communication both about developments that improve the translation of money into sound science and also how to turn scientific findings into meaningful action (management or policy). The urgent need to enhance the communication between conservation scientists and those who can implement management has been acknowledged (Memmott et al., 2010; Milner–Gulland et al., 2010). The geographical imbalance in opportunities for developing conservation science seems to have drawn less attention (but see McGowan, 2010a). There is much to gain by enhancing the communication between research cultures where there is a decades–long tradition of ecological study with those parts of the world that have much younger research traditions, but where numbers of species (especially those at risk of global extinction), is generally far higher. Nowhere are these opportunities for increasing communication more apparent than amongst those who study and manage Galliformes, widely known as the gamebirds. This avian Order contains some of the most intensively and extensively studied species in the world (such as grey partridge Perdix perdix, willow ptarmigan [= red grouse] Lagopus lagopus, bobwhite quail Colinus virginianus) as well as a very high proportion of threatened species (see below). This knowledge provides two exceptional opportunities to use lessons learnt from a relatively small number of very well studied and intensively managed species for much wider conservation benefit. The first of these is to use knowledge of conservation science of the better–studied species to enhance the survival prospects of the more poorly known species and, second, to accelerate the translation of science into management, especially for species where intervention is most needed to prevent population or species extinctions. Put simply, there is a significant body of information that is not being used for the widest possible benefit. The consequence of this is that extinctions (local or global) are more likely to result because appropriate knowledge is not being applied to address pressing challenges. Here we suggest how the gamebird science community can make a significant contribution to

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preventing species extinctions and improving the status of threatened species through better communication leading to wider application of appropriate approaches and techniques. We will do this by outlining the current global policy context, and by reviewing the threat status of Galliformes and the actions necessary to reduce their risk of extinction. Finally, we suggest ways in which scientific efficiency may be enhanced by indentifying key issues in understanding the status of the most threatened species and determining the context–specific action necessary. Global policy context Convention on Biological Diversity 2020 targets In 2002, the 193 Parties to the Convention on Biological Diversity (CBD) agreed to reduce the rate of biodiversity loss significantly by 2010 (CBD, 2002). The wider importance of biodiversity to human well– being was recognised by the adoption of this target contributing towards Millennium Development Goal 7: Environmental sustainability (UN, 2005). In early 2010, there were a variety of analyses that showed this target, vague as it was, had not been met. Most important was the CBD’s own assessment Global Biodiversity Outlook 3 (Secretariat of the Convention on Biological Diversity, 2010). Later that year, the Parties to the Convention adopted 20 new targets for biodiversity conservation that were much more specific and highly ambitious. The target for species was agreed as: “By 2020 the extinction of known threatened species has been prevented and their conservation status, particularly of those most in decline, has been improved and sustained”. (CBD, 2010) At the same time, the importance of Galliformes as wild relatives of significant food species received attention at the 6th Session of the Intergovernmental Technical Working Group on Animal Genetic Resources for Food and Agriculture (see McGowan, 2010b). The United Nations’ Food and Agriculture Organisation is charged with implementing the Global Action Plan for the Conservation of Animal Genetic Resources and although this emphasizes the need for action to conserve rare breeds and domesticated varieties, the importance of wild relatives is touched upon (FAO, 2007). There are clear linkages between the implementation of this plan and CBD (2010), in which target 13 states: 'By 2020, the genetic diversity of cultivated plants and farmed and domesticated animals and of wild relatives, including other socio– economically as well as culturally valuable species, is maintained, and strategies have been developed and implemented for minimizing genetic erosion and safeguarding their genetic diversity'. Three factors (wild relatives of species important to humans, long history of research and management in some species, and overall threat status), therefore, place Galliformes in a unique position to advance not only their own conservation, but as a model for increasing the efficient application of scientific and management developments to species most at risk.

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Northern Europe 7 Eastern Europe Western Europe 16 9 Southern Europe 9

Northern America 18

Eastern Asia 57 Western Asia South–central Asia 12 55

Northern Africa 12 Central America & Caribbean 36

Southern America 55

South–easthern Asia 72

Western Africa 13 Eastern Africa 35 Middle Africa 31 Southern Africa 16

Oceania 14

Fig. 1. Global distribution of the 290 Galliformes species. Fig. 1. Distribución mundial de las 290 especies de Galliformes.

Galliformes and their habitats

Threat status

It is obvious that a range of research, management and monitoring approaches and techniques will be needed in different circumstances. For example, assessing the status of difficult to detect species inhabiting lowland rainforest in Indonesia provides very different challenges compared with Himalayan forest species that are often seen and heard regularly. Therefore, when understanding how best to exploit the opportunities for learning from the better studied and more intensively managed species it is necessary to understand the diversity of habitats and ecological requirements of the more threatened species. For conservation purposes the most convenient list of species is that on the IUCN Red List and this describes 290 species of Galliformes (IUCN, 2011). The Order is made up of the following families: Phasianidae (181 species), Odontophoridae (31), Cracidae (51), Numididae (6) and Megapodidae (21) (BirdLife International, 2012a). They have a worldwide distribution (fig. 1) with concentrations of species in eastern and southern Asia and in South America. As the species occur in all major habitat types and from sea level to high mountains and from the equator to high latitudes (McGowan, 1994), the range of ecological contexts in which they occur is considerable. This has direct implications for the sort of survey, research and management approaches that can be applied, depending upon how readily species are detected, how easy it is to work in each habitat and terrain, and the nature of interactions with human communities.

The global index of species threat status is the IUCN Red List (www.iucnredlist.org) which documents the extinction risk of all assessed species (see Mace et al., 2008; Vié et al., 2009). In 2011,

Table 1. Number of Galliformes in each Red List category in 2011: N (toward extinction, from bottom to top); P (cumulative percentage, from top to bottom). Tabla 1. Número de Galliformes en cada categoría de la Lista Roja en 2011: N (hacia la extinción, de abajo a arriba); P (porcentage acumulado, de arriba a abajo). N

Category

Extinct since 1600

0.7%

Extinct in the wild

1.0%

Critically endangered

2.8%

Endangered

11%

Vulnerable

26%

Near–threatened

39%

177

Least concern

100%

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Hunting and trapping

Agriculture & livestock

Logging & harvesting

Natural modifications

Energy & mining works

Climate effects

Transport effects

Building developments

Invasive species, etc.

30 40 50 Number of species

Fig. 2. The threats facing Galliformes listed as threatened on the IUCN Red List. Fig. 2. Las amenazas a las que se enfrentan las Galliformes catalogadas como amenazadas en la Lista Roja de la UICN.

the 9,920 species of bird recognised by BirdLife International had been evaluated against the IUCN Red List criteria and 1,253 (12.5%) were considered threatened with extinction (BirdLife, 2012b). In contrast, 76 (26%) of the 290 species of Galliformes were included on the Red List. Table 1 shows that species do move towards extinction and that at present there is one species that only survives in captivity. The last wild record of the Alagoas curassow Mitu mitu, which inhabits the Atlantic Forest of Brazil, was in the late 1980s and it is now considered 'Extinct in the Wild'. A further example of a species that is moving closer to extinction is the uplisting of Edwards’s pheasant Lophura edwardsi from Central Vietnam from Endangered to Critically Endangered that will take place on the forthcoming 2012 IUCN Red List (BirdLife International, 2012c). Given the intense pressures and scarcity of resources it is critical to use the time, funds, expertise and people that are available to best possible effect. To achieve this, action therefore, falls within two extremes: broad–based policy interventions intended to address widespread issues and species–specific programmes designed to counter the particular pressures and constraints that are threatening individual species. Two points are clear: 1) some issues are best addressed at a policy level, and 2) the resources needed to develop and implement detailed conservation strategies for all 76 threatened Galliformes are beyond reasonable expectations.

Taking action Policy Assessment of the pressures listed for the threatened species shows that over–exploitation and habitat change are the overwhelming issues (fig. 2; derived from IUCN, 2011). It is clear that there is a need to promote policy to reduce overexploitation and the worst effects of habitat change at various political and administrative levels. As this involves the advocacy of Galliformes science rather than increasing its efficiency and quality, it is not considered further here. Species priorities Critically Endangered species In 2003, the World Pheasant Association reviewed the species that were then listed as Critically Endangered and concluded that there was little concerted action underway for three of them: Djibouti francolin Francolinus ochropectus, gorgeted wood quail Odontophorus strophium, and Trinidad piping– guan Pipile pipile. This led to a population survey of a key site for the wood quail in 2003 (Turner & Donegan, 2006) and the resulting population estimates, together with information from other sites, led to the gorgeted wood quail being downlisted to Endangered in 2008 (BirdLife International, 2012d). This left two Critically Endangered species requiring directed conservation action.

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Table 2. The 31 countries in which the 24 Endangered Galliformes occur. Tabla 2. Los 31 países en los que se encuentran las 24 Galliformes en peligro.

14 single country endemics Angola, Brazil, Cameroon, China (2 species), Colombia (2 spp.), Indonesia (2 spp.), Tanzania, Tonga, US, Vietnam (2 spp.) 10 remaining endangered species occur in 17 further countries Argentina, Bolivia, Cambodia, Congo DR, Ecuador, Guatemala, Lao PDR, Malaysia, Mexico, Myanmar, Northern Mariana Islands, Palau, Paraguay, Peru, Thailand, Uganda, Venezuela

Several activities have been undertaken to gather information and promote the conservation of both the Djibouti francolin and the Trinidad piping–guan and these led to the development of Species Conservation Strategies (see IUCN/SSC, 2008) in 2010. These strategies seek to bring together both those who can affect the species’ conservation status and those who may be affected by the resulting action (or lack of it). This group of stakeholders then develops a vision for the species and the practical goals, objectives and actions necessary to realise that vision (IUCN/SSC, 2008). These are resource intensive activities, as they involve bringing together a variety of people and require a significant amount of time if

the planning process is be comprehensive and thus allow the resulting strategy to stand the best chance of successful implementation. Success should, in due course, be measured by the downlisting of the target species and, ultimately, its removal from the Red List. Endangered species Critically Endangered species merit the most intensive attention as they are, by definition, those most at risk of extinction. This is more manageable because there are relatively few species in few countries. In contrast, there are 24 Endangered species spread across 31 countries (table 2, fig. 3),

Fig. 3. Distribution of single–country endemic Endangered Galliformes. Fig. 3. Distribución de las Galliformes en peligro, endémicas de un único país.

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making it a significant challenge to provide sufficient research, management and monitoring effort for these species not only because they are widely dispersed, but also because they are found in countries where capacity is often limited compared with conservation science needs. These are the species, therefore, where increased efficiencies have the biggest potential to contribute towards averting species extinctions. The first step is to ensure that the species is appropriately categorised on the Red List. The five criteria against which each species is assessed are (IUCN, 2001): (i) Reduction in population size ≥ 70% over the last 10 years or 3 generations whichever in the shorter; (ii) Geographic range small (Extent of Occurrence of < 5,000 km2 or Area of Occupancy < 500 km2) and fragmented, declining and extreme fluctuations; (iii) Population size < 2,500 mature individuals and declining; (iv) Population size < 250 mature individuals; and; (v) Quantitative analysis showing the probability of extinction in the wild is at least 20% within 20 years or five generations. Eight species are currently listed under Criterion A, 14 under B, 13 under C and one under D. None of these species are listed as a result of an acceptable quantitative population viability analyses. As species should be listed under all criteria that they meet, the total above (36) is greater than the number of species (24). Science Assessing status Recent advances in methods of assessing population sizes (Buckland et al., 1993; MacKenzie et al., 2002), geographic ranges sizes (Phillips et al., 2006) and in viability analyses (Lacy, 2000; Akçakaya & Root, 2002) offer considerable potential for generating appropriate and reliable data on poorly known Endangered species. Gamebird ecologists have a significant opportunity to contribute because some of these techniques have been very widely applied to a few highly studied galliform birds. For example, methodological arguments about how, when and what index to use when counting Galliformes have already been largely resolved in North America and Europe (Warren & Baines, 2011; Willebrand et al., 2011; Calladine et al., 2009). Distribution modelling (for example, Aldridge et al., 2012; Graf et al., 2009; Gottschalk et al., 2007) and population viability analysis (Lu & Sun, 2011; Johnson & Braun, 1999; LaMontagne et al., 2002) have become key to assessments of gamebird populations. This practical experience and an understanding the biological requirements of the species suggest that lessons are being learnt which can now be applied to Endangered species, most of which have been subject to little or no quantitative field study. Insights generated by using these methods in field studies of Endangered species can be combined with remotely gathered data to produce

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powerful approaches to understanding species status in remote and challenging habitats. Even where location data are scarce, techniques such as Resource Selection Functions (Boyce et al., 2002) combined with high quality satellite images can produce predicted distribution maps which in turn allow the selection of priority areas and the efficient targeting of survey effort (e.g. Gottschalk et al., 2007). All of this would allow better understanding of two key issues. Firstly, do we have reasonable assessments of extinction–risk for each species? If we do, this would ensure that effort is targeted where it is most needed and the downlisting of the gorgeted wood–quail as a result of new knowledge is an example of this. Secondly, generating quantitative data for the parameters used to determine extinction risk will help to determine the factors that have led to species being considered to have a high risk of extinction. Understanding threats IUCN has developed a standardised classification for threats to species (IUCN, 2012a) and all new assessments specify threats according to these schemes. Although a wide range of threats have been documented for Endangered Galliformes, three stand out because of the number of species that they affect: agriculture involving annual and perennial non–timber crops, hunting and trapping, and logging and wood harvesting (fig. 4). There is extensive literature from intensively managed species, typically from Europe and North America, which explores these threats, both in terms of their impacts on particular galliform species and the actions that are necessary to mitigate those impacts (Bunnefeld et al., 2011; Dallimer et al., 2010; Pearce & Higgins et al., 2007). Conversion of habitat to intensive agriculture has been responsible for declines in species such as red grouse and black grouse in Europe (Patthey et al., 2012; Dallimer et al., 2010; Ludwig et al., 2009) and prairie grouse in North America (Riley, 2004). Habitat reclamation and management such as set aside schemes offer potential to halt these declines (Riley, 2004; Patthey et al., 2012). Sustainable harvesting of gamebirds has a long established tradition in Europe and North America. Management techniques that have been applied to these species successfully could be applied to threatened species elsewhere that are hunted in unsustainable numbers. These techniques include bag limits (Sandercock et al., 2011), habitat management (Patthey et al., 2012) and predator control (Summers et al., 2004). The effects of management practices on gamebird populations by timber harvesting in North America and Europe are well known. The effects of logging rotations, remnant forest strips and fragmentation have all been addressed (for example, Potvin & Courtois, 2006; Giroux et al., 2007; Pearce–Higgins et al., 2007; Borchtchevski et al., 2009) leading to robust and testable habitat management recommendations.

Number of Galliformes

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26 24 22 20 18 16 14 12 10 8 6 4 2 0

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Hunting & trapping Annual and perennial NTCs Logging & wood harvesting

1.1 1.2 2.1 2.2 2.3 3.1 3.2 4.1 4.2 5.1 5.2 5.3 6.1 6.2 6.3 7.1 7.2 8.1 10.1 11.1 11.2 11.3

5 Threats

Fig. 4. Threats to Endangered Galliformes categorised by IUCN Red List Threats Classification Scheme: 1. Residencial and commercial development; 2. Agriculture and aquaculture; 3. Energy production and mining; 4. Transportation and service corridors; 5. Biological resource use; 6. Human intrusions and disturbance; 7. Natural system modifications; 8. Invasive and similar; 10. Geological events; 11. Climate change and severe weather. Fig. 4. Amenazas para las Galliformes en peligro catalogadas según el sistema de clasificación de amenazas de la Lista Roja de la UICN: 1. Desarrollo residencial y comercial; 2. Agricultura y acuicultura; 3. Producción de energía y minería; 4. Transporte y servicio de corredores de transporte; 5. Uso de los recursos biológicos; 6. Intrusiones y perturbaciones humanas; 7. Modificaciones naturales del sistema; 8. Invasoras y similares; 10. Eventos geológicos; 11. El cambio climático y clima adverso.

Underpinning action

Discussion

Understanding the probable consequences of action and the nature of threats are pre–requisites for defining as precisely as possible the action to be undertaken to mitigate threats. At present, the actions proposed for Endangered Galliformes are primarily site–based (fig. 5). In many cases, however, these proposals, classified under the IUCN Actions Classification Scheme (IUCN, 2012b) are based on very limited information and will benefit significantly from both better knowledge of the species as described above and comparison with management actions that have been tried and are documented in other galliform species. Some Galliformes species have been subject to relatively well–researched and documented interventions and these may provide powerful lessons for poorly known Endangered species in capacity–limited areas of the world. Already, habitat fragmentation, a dominant paradigm in bird ecology, is becoming a major focus in the study of Galliformes in capacity– limited areas of the world (e.g. Cabot’s tragopan in China, Deng & Zheng, 2004; chestnut–breasted hill– partridge in Indonesia, Nijman 2003). The application of these same techniques to Endangered galliform species must now be the next step.

The 290 species of Galliformes are a remarkable group of birds. As a whole they are very important to humans and contain some of the most studied species in the world, but for many reasons they are highly threatened. These factors provide a valuable opportunity to both translate lessons from well studied to poorly studied species, especially those most at risk of extinction, and also in turning science into effective management. As global policy provides a context that explicitly promotes threatened species conservation, this offers an additional incentive to enhance the efficiency of research, management and monitoring for the most threatened species. The wealth of applied research that has been conducted on partridges, quail, pheasants and grouse in Europe and North America has provided extensive literature that can be drawn on by those working on the most threatened species. Areas where such research offers especial insights include methods of population estimation (Warren & Baines, 2011), habitat use assessment (Dzialak et al., 2011), breeding ecology and success (Draycott et al., 2008; Kurki et al., 2000) and measuring and understanding mortality (Stephenson et al., 2011). Management has been explored in a wide variety of contexts and

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Number of Galliformes

18 16 14 12 10

8 6 4 2 0

2 3 4 5 Conservation action for endangered Galliformes

Fig. 5. Actions proposed for Endangered Galliformes categorised using the IUCN Red List Actions Classification Scheme: 1. Land/water protection; 2. Land/water management; 3. Species management; 4. Education and awareness; 5. Law and policy; 6. Livelihood, economic and other incentives. Fig. 5. Medidas propuestas para las Galliformes en peligro catalogadas según el sistema de clasificación de medidas de la Lista Roja de la UICN: 1. Protección tierra/agua; 2. Gestión tierra/agua; 3. Gestión de las especies, 4. Educación y concienciación; 5. Legislación y política; 6. Sustento, incentivos económicos y de otro tipo.

although there are clear insights to be gained, these will be largely in understanding the ecological issues that management is designed to tackle, rather than socio–economic context in which each intervention will take place. For example, there are clear lessons to be drawn from past reintroduction efforts (World Pheasant Association and IUCN/SSC Re–introduction Specialist Group, 2009) in terms of numbers for release strategy, but where recovery programmes are community based, the involvement of the local community where the project will take place in developing and implementing the project is a key factor in its success (Waylen et al., 2010). Specific needs There are several areas that are immediate priorities for exploring what lessons can be learnt from combining knowledge from the well–understood species with developments in conservation science approaches (such as assessing the impact of off–take) and applying them to species with the most pressing needs (i.e. those that are Critically Endangered and Endangered). A key constraint in promoting knowledge and action for the most threatened species is simply encouraging researchers and conservationists to get out into the field to undertake status assessments and determine what needs to be done. Conducting basic status assessments often seems very daunting for researchers in countries where there is limited capacity and little tradition of ornithological field work because the effort that must be expended in order to gather sufficient data on species can be perceived to be considerable. Therefore, clearer

guidance on field techniques and their application in various circumstances (habitat, topography and species detectability) would make a significant contribution in both demonstrating that gathering meaningful data is possible and outlining how it might be done. New methods for population estimation are typically developed in countries that are relatively resource–rich and which have easier access across study areas and explicit efforts to test these on some of the most threatened species that inhabit areas where research is logistically more difficult would contribute significantly to Galliformes conservation. If this can be combined with greater support for new researchers through the mentoring of individuals or small groups, the increase in field effort on the most threatened species would be marked. The main threat that makes Galliformes more highly threatened than most other avian Orders is over–exploitation. At present this is based on a perception that observations of snaring or wild meat extraction are indicative of removal outstripping recruitment. Whilst this may be suitably precautionary, further information on the scale of exploitation, allied to a better understanding of the genetic and demographic consequences of exploitation on hunted populations with various ecological characteristics (large population sizes, geographically restricted, habitat specialists etc). This is an area where significant contextual lessons could be learnt from the well–studied species to those that are poorly known and/or considered to be threatened with global extinction. Finally, learning lessons about successful interventions may result in improving the efficiency of management for the most threatened species. Interventions

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have both an ecological and social component, with the latter varying from place to place. There are now sufficient assessments of management interventions (some formally written up, others not) from the well– studied galliform species that general principles should be explored. It may lead to important conclusions about basic characteristics of management that may have especially positive impacts, such as stage of life history or reproductive cycle impacted or the needs of suites of particular species. Such findings may help develop more effective management approaches and methods for the most threatened species with greater efficiency. As a way of approaching the analysis of a species conservation status and working to achieve its conservation, an Adaptive Resource Management (ARM) approach (Walters, 1986) offers a potential model to link scientific research and management outcomes for the benefit of Galliformes conservation (Conroy & Peterson, 2006). ARM acknowledges uncertainty inherent in ecological data collection and allows for managers to learn about ecosystems at the same time as managing them (Lancia et al., 1996). This approach is being used to manage populations of well–known Galliformes, particularly in North America (e.g. sage grouse in Canada, Canadian Sage Grouse Recovery Team, 2001). More formal incorporation of ARM into other areas of gamebird conservation could significantly enhance our ability to meet the CBD 2020 target for species conservation. References Akçakaya, H. R. & Root, W., 2002. RAMAS Metapop: viability analyses for stage–structured metapopulations (version 4.0). Applied Biomathematics, Setauket, NY, USA. Aldridge, C. L., Saher, D. J., Childers, T. M., Stahlnecker, K. E. & Bowen, Z. H., 2012. Crucial nesting habitat for Gunnison sage–grouse: A spatially explicit hierarchical approach. Journal of Wildlife Management, 76: 391–406. Barnosky, A. D., Matzke, N., Tomiya, S., Wogan, G. O. U., Swartz, B., Quental, T. B., Marshall, C., McGuire, J. L., Lindsey, E. L., Maguire, K. C., Mersey, B. & Ferrer, E. A., 2011. Has the Earth’s sixth mass extinction already arrived? Nature, 471: 51–57. BirdLife International, 2012a. Data Zone. Available at www.birdlife.org/datazone/species/search. Accessed on 15 March 2012. – 2012b. Birds on the IUCN Red List. www.birdlife. org/action/science/species/global_species_programme/red_list.html Accessed 15 March 2012. – 2012c. Final decisions for the 212 IUCN Red List. Available at www.birdlife.org/globally–threatened– bird–forums. Accessed 15 March 2012. – 2012d. Gorgeted wood–quail Odontophorus strophium. www.birdlife.org/datazone/speciesfactsheet. php?id=333. Accessed 15 March 2012. Borchtchevski, V. G., Hjeljord, O., Wegge, P. & Sivkov, A. V., 2003. Does fragmentation by logging reduce grouse reproductive success in boreal forests?

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Post–breeding movements and migration patterns of western populations of common quail (Coturnix coturnix): from knowledge to hunting management J. D. Rodríguez–Teijeiro, F. Sardà–Palomera & M. Puigcerver Rodríguez–Teijeiro, J. D., Sardà–Palomera, F. & Puigcerver, M., 2012. Post–breeding movements and migration patterns of western populations of common quail (Coturnix coturnix): from knowledge to hunting management. Animal Biodiversity and Conservation, 35.2: 333–342. Abstract Post–breeding movements and migration patterns of western populations of common quail (Coturnix coturnix): from knowledge to hunting management.— We investigated the patterns of post–breeding movements of the common quail (Coturnix coturnix) in the Iberian peninsula with the aim of describing its migratory phenology and some physiological features of individuals. This information is needed to adjust hunting seasons in an optimal way. We worked with two data–sets: a) captures made in a non–breeding site (Garraf) from August to October in 2009 and 2010; b) post–breeding recoveries of individuals ringed in Europe and recaptured in Spain between 1933 and 2005. The results showed that post–breeding movements in Garraf occur in two waves: a first wave that occurs around 10 VIII and is mainly composed of non–sexually active yearlings that do not correspond physiologically to migrants, and a second much more intense wave, which occurs around 17 IX and is mainly composed of non–sexually active migrant yearlings. The hunting season in Spain takes place mainly during the first wave, preserving the passage of migrant individuals from Spain and other European countries. Information on the post–breeding movements in other Spanish regions and other European countries where the common quail is a popular game species would improve timing between the hunting season and migration by providing more precise recommendations for hunting management. Key words: Hunting season, Game management, Ring recovery, Migration phenology, Iberian peninsula. Resumen Patrones de movimientos y de migración postcría en la población occidental de codorniz común (Coturnix coturnix): algunas recomendaciones de gestión cinegética.— Hemos investigado los patrones de los movimientos postcría de la codorniz común (Coturnix coturnix) en la península ibérica con el fin de describir su fenología de paso migratorio y algunas características fisiológicas de los individuos. Esta información es necesaria para un ajuste óptimo de los períodos de caza. Hemos trabajado a partir de dos conjuntos de datos: a) capturas efectuadas en una zona que no es de cría (Garraf) de agosto a octubre en 2009 y 2010; b) recuperaciones, posteriores a la presunta época de cría, de individuos anillados en Europa y recapturados en España durante el período 1933–2005. Los resultados obtenidos muestran que los movimientos postcría en Garraf están formados por dos oleadas: una primera, que se produce sobre el 10 VIII, formada principalmente por jóvenes del año inactivos sexualmente que no son fisiológicamente migrantes; y una segunda, mucho más intensa, que se produce sobre el 17 IX, formada principalmente por migrantes jóvenes del año inactivos sexualmente. La época de caza en España tiene lugar principalmente durante la primera oleada, preservando el paso de los migrantes provenientes de España y de otros países europeos. La información de los movimientos postcría en otras regiones españolas y en otros países europeos en los que la codorniz común es una especie cinegética popular, permitiría mejorar el ajuste entre el período de caza y la migración, proporcionando recomendaciones de gestión cinegética más precisas para esta especie. Palabras clave: Media veda, Gestión cinegética, Recuperaciones de anillas, Fenología de la migración, Península ibérica.

ISSN: 1578–665X

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Received: 30 I 12; Conditional acceptance: 27 II 12; Final acceptance: 10 IV 12 J. D. Rodríguez–Teijeiro, Dept. de Biologia Animal, Fac. de Biologia, Univ. de Barcelona, Avda. Diagonal 645, 08028 Barcelona, Espanya (Spain).– F. Sardà–Palomera, Centre Tecnològic Forestal de Catalunya, Àrea de Biodiversitat, carretera Sant Llorenç de Morunys, km 2 (ctra. vella), 25280 Solsona, Espanya (Spain).– M. Puigcerver, Dept. de Didàctica de les Ciències Experimentals i la Matemàtica, Fac. de Formació del Professorat, Univ. de Barcelona, Psg. Vall d’Hebron 171, 08035 Barcelona, Espanya (Spain).

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Introduction The common quail (Coturnix coturnix) is a very abundant and widespread migratory galliform species whose distribution area encompasses the Western Palaearctic, and Western and Central Asia (Gallego et al., 1997). Its estimated population ranges from 35,000,000 to 300,000,000 individuals, with an extent of occurrence of 21,300,000 km2 (Birdlife International, 2004). In Europe, a range of 2,800,000–4,700,000 breeding pairs has been estimated and, according to Gallego et al. (1997), 33–57% of the European Union pairs breed in Spain. Currently, the species is considered to have an unfavourable conservation status in Europe (SPEC 3), with depleted populations and a large historical decline (Burfield, 2004). In spite of this, the species is also considered a game bird in many countries within its distribution range (Guyomarc’h, 2003). In Spain, in particular, it is a very popular game species, with a mean annual hunting bag of 1,381,503 individuals (standard deviation: 268,812; data calculated from the hunting bags of 1973–2008, collected from the Yearbook of Agro–alimentary Statistics of the Spanish Ministry of Agriculture, Fishing and Food). Moreover, 33–57% of the quails of the European Union breed in Spain (Perennou, 2009), a figure that is not surprising as Spain is a major pass way for individuals that are migrating to Africa from several Western European countries (unpub. data). Therefore, data on both breeding pairs and hunting bags strongly suggest that management of the species and conservation measures that are adopted in Spain are extremely important for the entire quail population in Europe. According to the European Union Management Plan for common quail for the period 2009–2011 (Perennou, 2009) and the European Birds Directive 2009 147/EC, hunting periods in EU member states should concord with information on the breeding period, and hunting activity should not affect late breeding birds or birds during spring migration. However, it is also necessary to know the post–breeding movements and the post–breeding migratory patterns to correctly adjust the hunting period. These movements and patterns are generally very poorly understood and more specifically, information concerning the Iberian Peninsula is extremely scarce (see however Guyomarc’h et al., 1989; Rodríguez–Teijeiro et al., 1996). The common quail shows not only migratory movements, but also nomadic movements (Sinclair, 1984) during the breeding season in search of suitable but ephemeral habitats, mainly winter cereal crops (Rodríguez–Teijeiro et al., 2009). These movements can be divided into latitudinal (or aestival) movements from northern Africa to Europe (Munteanu & Maties, 1974) and elevation (or transhumant) movements within Europe (Davis et al., 1966, Heim de Balsac & Mayaud, 1962; Puigcerver et al., 1989). Movements of males in search of females throughout the breeding season have also been described (Rodríguez–Teijeiro et al., 2006). These movements are not sporadic, but are part of the annual cycle of the species and, as Wernham et al. (2002) suggest, they seem to be

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firmly set to maximize the production of yearlings, in a remarkable sequential breeding strategy similar to that of some butterflies and moths. This extremely high mobility of the species makes it more complicated to determine the current suitability of hunting periods and, not surprisingly, there are no data available concerning how the post–breeding movements and migration might be affected by the traditional hunting calendar. This study aimed to provide new data concerning patterns of post–breeding movements in the Iberian peninsula and also to determine how the hunting season is related to these movements. The evaluation of how this could affect European quail populations would hopefully lead to the development of new management recommendations. Material and methods Data collection The data presented here were collected from two sources, allowing two different approaches: (1) The capture and ringing of post–breeding moving individuals using mist–nets and electronic decoys in a non– breeding area in Garraf Natural Park, north–eastern Spain; this source of data provided information on the temporal migration patterns of the common quail at a local scale. (2) Quail ringing recovery data in Spain; analysis of the post–breeding recoveries of individuals ringed in Europe and recaptured in Spain during the period 1933–2005 provided information on the temporal migration patterns of the species at a large scale. Captures of post–breeding moving individuals The method used for collecting data on post–breeding individuals involved the use of mist nets and call playback in a non–breeding area. Fieldwork was carried out in Garraf Natural Park in the north–eastern coast of Spain. This area is characterised by rough landscape with typical Mediterranean sclerophyllous vegetation consisting mainly of shrubs and pine trees. The Garraf Natural Park is not a suitable place for breeding for the common quail (Rodríguez–Teijeiro et al., 2004). We selected a clear flat area at about 350 m from the edge of a 190 m cliff that drops to the sea (41º 15' N, 1º 52.7' E). During 2009 and 2010, captures were carried out approximately once a week from 15 VII to the end of October, when quail migration finished in this area. A total of six mist nets, each one with 6 bags of 25 mm mesh, 12 m long and 3 m height, were set up in the study area in 2009, and 18 mist–nets with the same features were set up in 2010. The nets were placed around a digital, quail call playback device which could be heard within approximately a 2 km radius in optimal conditions. The call was played after sunset from 22:00 h at night to 8:00 h in the morning to attract post–breeding moving quails towards the nets. The use of electronic decoys may result in some bias (see for example Weatherhead & Greenwood,

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1981). For this reason, we checked that the sample of individuals captured in Garraf (presumably migrant individuals coming from northern areas) does not show bias in sex and age composition when compared to a sample of 302 individuals of known sex hunted during the same period in northern Spain (Llivia) and in France (Ariège and Cavalerie) and with a sample of 288 individuals of known age hunted in the same places. No statistical differences were found either in sex composition (c12 = 0.871; p = 0.35) or in age composition (c12 = 1.80; p = 0.18), thus indicating that no significant bias of sex and age composition was linked to the use of electronic decoys in the studied sample. Quails captured in the nets were immediately collected, ringed, measured and released. Information about sex and age (following Saint–Jalme & Guyomarc’h, 1995) was recorded; individuals classified as EURING 3 age code are named hereafter yearlings, whereas individuals classified as EURING 5 and EURING 6 are named hereafter adults, due to the extreme sexual precocity of the common quail (Guyomarc’h, 2003). Measurements of the width of the pectoral lipid band, a good indicator of the migratory impulse according to Guyomarc’h & Belhamra (1998),were taken for all captured quails. Finally, the sexual development of the captured quails was assessed by the detection of presence or absence of the proctodeal foam (Seiwert & Adkins–Regan, 1998) and by measuring the length of the cloacal vent in a sample of 319 individuals; less than 4.5 mm would indicate that individuals are not sexually active (Fontoura et al., 2000; Guyomarc’h & Belhamra, 1998). Quail ring recoveries in Spain Data from ring recoveries in the Iberian peninsula of quails ringed in Europe (from 1933 until 2005) were provided by EURING Data Bank. Most quail ring recoveries come from hunted birds. Quail hunting is a widespread tradition in the Iberian peninsula; the Spanish hunting law ('Ley de caza') regulating the hunting of migrating birds was enacted in 1902 and it stated that migrating birds could be hunted during any season of the year. This law was modified in 1970 to establish a hunting period for migrating birds (Streptopelia turtur, Columba palumbus and Coturnix coturnix) from 15 VIII to 15 IX; there may be small variations between years and regions. The general hunting season for all game species (including the common quail) opens again from 12 X until mid–February. Every ring recovery has associated information about the ringing date in the usual format, which we transformed into days of the year (days elapsed since 1 I), the ringing location, and the recovery date and location. Unfortunately, information about the sex and age of the individuals is mostly incomplete. For the analysis, we only selected rings recovered in the Iberian peninsula within the same year of ringing. To ensure that wintering individuals were not included, all the recoveries subsequent to 30 X were excluded from the analysis. These data were also divided in three groups: Spanish recoveries (quails ringed in Spain), inter-

national (quails ringed elsewhere in Europe) coastal recoveries (recoveries located in the Mediterranean slope, which is an area at less than 50 km from the shoreline and limited by Mediterranean coastal mountain chains), and international inland recoveries (the remaining recoveries). This division was made to explore the possible differences in quail migration patterns suggested in other studies (Rodríguez–Teijeiro et al., 2009; Zduniak & Yosef, 2008; Zuckerbrot et al., 1980). Statistical analysis We applied parametric tests (multi–way ANOVA, Student t–test, Chi–square test) and when conditions of application were not fulfilled, non–parametric tests were applied (Mann–Whitney U–test, median test and Fisher exact test). In this case, descriptive statistics of central trend and dispersion used were the median and quartiles, respectively. PASSW and Statistica software were used for calculations. Results Quail captures in Garraf A total of 530 quails (85.10% were yearlings) were trapped in Garraf during the study periods in 2009 and 2010. Surprisingly, the number of quails trapped was higher in 2009 (N = 275) than in 2010 (N = 255), when the number of mist–nets was tripled. Captured birds showed that there are two different movement waves in the post–breeding passage (figs. 1, 2). In both years, the first wave occurred in August (median: 13 VIII 2009, quartiles 25–75: 219–224 days, and 7 VIII 2010, quartiles 25–75: 217.25–219 days), whereas the second , more intense wave occurred in September and October (median: 13 IX 2009, quartiles 25–75: 254–263 days, and 22 IX 2010, quartiles 25–75: 255–276 days), this latter difference being statistically significant (Mann–Whitney U–test Z = 7.26; N1 = 280; N2 = 269; p < 0.01). No significant differences were found in sex composition of the first and second wave when analyzing them by age classes in the two years (Yearlings: c12 < 1.84; p > 0.33; Adults: Fisher exact test; p > 0.39), or in the proportion of ages between waves in the two years (c12 < 0.43; p > 0.51). No differences were found in the cloacal vent between yearling and adult males in the two years of study (two– way ANOVA, year factor: F(1,315) = 0.036; p = 0.88; age factor: F(1,315) = 4.14; p = 0.29); interaction year x age: F(1,315) = 0.339; p = 0.56). However, individuals belonging to the first wave had a higher cloacal vent (mean ± SE = 5.23 ± 0.17, n = 54) than those belonging to the second wave (mean ± SE = 4.39 ± 0.05, n = 262). None of the individuals in either wave showed proctodeal foam, indicating they were not sexually active (Guyomarc’h et al., 2001; Seiwert & Adkins–Regan, 1998). In 2010, yearling females showed a delay in migration (fig. 2) when compared to yearling males (median day of yearling males: 266; quartiles 25–75%: 258–272; median day of yearling females: 272; quartiles 25–75:

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2009

50 40

Number of individuals

30 20 10 0 30

2010

25 20 15 10 5

0 190

210

230 250 270 Day of the year

290

310

Fig. 1. Number of individuals captured from August to October in 2009 and 2010 in Garraf. The grey band shows the hunting season for migrant species in Barcelona province (15 VIII to 4 IX). Fig. 1. Número de individuos capturados de agosto a octubre en 2009 y 2010 en Garraf. Las bandas grises corresponden a la época de caza de las especies migratorias en la provincia de Barcelona (15 VIII a 4 IX).

258–284; U–test Mann–Whitney Z = 2.49; N1 = 135; N2 = 68; p = 0.01). This phenological difference was, however, less clear in 2009 (median day of yearling males: 257; quartiles 25–75: 254–265; median day of yearling females: 255; quartiles 25–75: 253–262; Mann–Whitney U–test, Z = 1.78; N1 = 154; N2 = 97; p = 0.08). No differences were observed in the phenology of migration in adult individuals either in 2009 or in 2010 (Mann–Whitney U–test, U < 189; p > 0.40). From a physiological point of view, the width of the lipid band in yearlings showed differences between 2009 and 2010, being wider in 2009 (three–way ANOVA, year factor: F(1,436) = 18.66; p < 0.01). Differences were also found between waves, as individuals captured in the first wave showed a smaller lipid band than those captured in the second wave (wave factor: F(1,436) = 75.65; p < 0.001; fig. 3). However, no differences were found between sexes (factor sex: F(1,436) = 3.27; p = 0.07). Only the interaction between year and sex was significant; while males did not differ between years, females were less fat loaded in 2010 (interaction season x sex: F(1,436) = 6.59; p = 0.01; fig. 3). The width of the lipid band in adults showed non–significant differences between years (three–way ANOVA, year factor: F(1,67) = 30.67; p = 0.08). However, significant differences appeared between waves (wave factor:

F(1,67) = 33.49; p < 0.001). As in yearlings, there were non–significant differences between sex (sex factor: F(1,67) = 0.46; p = 0.50), and there was no interaction between factors (p > 0.05; fig. 3). Ring recoveries The distribution of ring recoveries in Spain showed that Spanish recoveries were captured very early (median: 20 VIII quartiles 25–75: 227–240, fig. 4). Of a total of 189 international recoveries, 55 (29.1%) were coastal recoveries, whereas 134 (70.9%) were inland recoveries. In the former, 52 (94.5%) were obtained before 1970, before the law was amended to establish a hunting season for migrating bird species; in the second case, 102 birds (76.1%) were recovered before this date. International recoveries show a first passage which mainly occurs inland, and a second passage which occurs along the Mediterranean coast (fig. 4). There were significant differences (fig. 4) in the median day of capture in the three groups of recoveries analysed (Spanish recoveries, international coastal recoveries and international inland recoveries), together with the median day on which individuals of the second wave were captured in Garraf (Median test:

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2009 6 5

Adult females

Adult males

Number of individuals

3 2 1 0 180 190 200 210 220 230 240 250 260 270 280 290 300 Day of the year 30

25 20

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15 10 5 0 180 190 200 210 220 230 240 250 260 270 280 290 300 Day of the year 2010 6 5 4

Adult females Adult males

Number of individuals

3 2 1 0 180 190 200 210 220 230 240 250 260 270 280 290 300 310 320 Day of the year 30 25 20

Yearling females Yearling males

15 10 5 0 180 190 200 210 220 230 240 250 260 270 280 290 300 310 320 Day of the year Fig. 2. Temporal migration patterns by sexes and ages in 2009 and 2010 in Garraf. Fig. 2. Patrones de migración temporal por sexos y edades en 2009 y 2010 en Garraf.

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Lipidic band width (mm)

Yearlings

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8 6

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5 3

2010

14 12

8 4

2009

Females Males

7 1st wave 2nd wave

2 3 2009 2010 Year

Females Males 2009

Year

2010

Fig. 3. Lipid bands of migrant quails in 2009 and 2010 in Garraf. Results on the left show whether the individuals belong to the first or second movement wave. Results on the right show interactions between year and sex factors. Mean and 95% confidence intervals are shown for yearling and adult males and females. Fig. 3. Bandas lipídicas de codornices migrantes en 2009 y 2010 en Garraf. Los resultados de la izquierda muestran según si los individuos pertenecen a la primera o a la segunda ola migratoria. Los resultados de la derecha muestran las interacciones entre los factores año y sexo. Se muestran la media y los intervalos del 95% de confianza para jóvenes del año y machos y hembras adultos.

c2685 = 596.75; p < 0.01). Moreover, a multiple post hoc comparison test showed differences (p < 0.01) in all the two–to–two comparisons with the exception of the comparison between the median day of capture of international coastal recoveries with regard to the median day on which individuals of the second wave were captured in Garraf (p > 0.01). Discussion The common quail post–breeding movement patterns through the Iberian peninsula were completely unknown to date. Our results in Garraf Natural Park, in the northeast of Iberian Peninsula, clearly suggest that an important movement flow occurs along the Mediterranean coast at latitude of 41ºN. Furthermore, they show that these movements are formed by two marked movement waves.

The first wave, which lasts from July to August (with a modal value of 13 VIII), is mainly composed of yearlings at the end of their sexually–active period (on the basis of their cloacal vent over 4.5 mm but, on the other hand, without proctodeal foam. According to their fat deposits, they cannot be considered physiologically as migrant individuals and should be considered more nomadic than migratory. This movement would belong, thus, to the movement patterns of the species during the breeding season (Munteanu & Maties, 1974; Puigcerver et al., 1989; Rodríguez–Teijeiro et al., 2006).The common quail is a farmland bird whose life cycle is mainly linked to winter cereal crops (wheat and barley mainly) and dense grassland (Guyomarc’h, 2003). However, during the breeding season, the species suffers massive habitat loss due to cereal harvesting. Thus, there is a radical landscape change in spring to the post–breeding migration passage, which induces common quail birds to move during the breeding season (Rodríguez–Tei-

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270

Day of the year

265 260

15 th September

255 250 245 Mean ± SE ± 1.96*SE

240 235

230

A B C D Phenology of quail ring recoveries

15th August

Fig. 4. Phenology of quail ring recoveries in Spain according to the ringing origin and recovery location. The hunting season for migrant species in Spain (15 VIII to 15 IX) after 1970 is shown: A. Spanish recoveries; B. International recoveries inland; C. International recoveries on the coast; D. Garraf second wave. Fig. 4. Fenología de la recuperación de anillas en España, según el lugar de origen del anillamiento y la localidad de recuperación. Se muestra la época de caza de las especies migratorias en España (15 VIII a 15 IX) después de 1970: A. Recuperaciones españolas; B. Recuperaciones internacionales en el interior; C. Recuperaciones internacionales en la costa; D. Segunda oleada en Garraf.

jeiro et al., 2009) from south to north and from lower areas (harvested in June) to higher areas (harvested in August). Thus, at the end of the breeding season, quail population movements in the Iberian Peninsula present a complex scenario, and quails that breed in the Peninsula show nomadic movements in search of remaining suitable habitats to match habitat loss caused by cereal harvesting (Gallego et al., 1993; Puigcerver et al., 1989; Rodríguez–Teijeiro et al., 2009), rather than a true migratory movement (Rodríguez–Teijeiro et al., 2006). The second wave is formed mainly of non–sexually active yearlings and occurs on average five weeks later. These birds have a cloacal vent under 4.5 mm, do not present proctodeal foam, and could be physiologically considered as migrants because of their lipid band. The passage of these common quails captured in Garraf during the second wave has a very similar phenology to that of international quails that were ringed and recovered on the Mediterranean coast (fig. 4) in a totally independent way. As the breeding cycle of the species in the Spanish strip between the Pyrenees and the Garraf Natural Park has finished by the time harvesting is over, and as the maximum passage is recorded one month later in Northeast Spain, the individuals belonging to this migratory wave must come mainly from the North of Europe. This post–breeding migration passage has also been reported in Egypt (Zuckerbrot et al., 1980), whereas in Italy and Israel an important spring passage has been described, with the post–breeding flow being much less intense (Macchio et al., 1999; Zduniak & Yosef, 2008).

In the two years of the study, age and sex composition of the first and second wave remained fairly constant. However, the phenology of yearling females suffered a delay in the second year. Furthermore, yearling females in 2010 also showed less developed fat deposits, suggesting that they were physiologically less prepared for the migratory passage. As females invest alone in brood care, any delay in the breeding attempt would be reflected in the migratory condition of the female fraction of the population. On the other hand, the proportion of adults showed constancy in their migratory condition over the years of study. Our results on the basis of the two years of study (which were very similar in terms of meteorological conditions) indicate that there is a set of nomadic movements at the end of the breeding season that coincides with the hunting season for migrant species in Spain. Moreover, it almost entirely respects the main migratory passage constituted by the coastal passage, probably affecting only individuals that breed in Spain and the international inland recoveries (fig. 1).This information should help to adjust timing between the hunting period and migration in coming years. However, as meteorological variability from year to year could affect the movements of individuals, data need to be collected over more years to clarify how the different meteorological conditions affect quail movements. This would provide more precise information and would permit reliable recommendations for adjusting hunting seasons to quail population movements. However, no efforts have been carried out to date in other Spanish regions, or in other European countries

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where the common quail is a popular game species, to study and describe its post–breeding migration patterns, which could vary from one country to another. Besides, the variability between years that could appear as a consequence of changes in meteorological conditions could have some influence on the movements of the individuals. Monitoring of post–breeding movements and post–breeding migration should thus be extended to all the regions and countries where the common quail is a game species. Furthermore, it should be carried out over years with varying meteorological conditions in order to improve adjustment between the hunting period and migration, thus complementing other measures described in Perennou (2009).Based on these findings we strongly suggest studies of this type should be conducted in other countries in coming years in order to gain further knowledge of the species and improve its management. Acknowledgements The authors are most grateful to the Catalan 'Direcció General de la Recerca', which provided financial support (Project nº 2009SGR481). The study is part of a Sustainable Hunting Programme promoted by José Mari Usarraga from the Hunting Federation of the Basque Country, which also provided financial support. We would like to thank the owner of the study area, Fidel Granada, for allowing us to install the capture station, the 'Diputació de Barcelona', and the secretary of the hunting estate, Josep Feliu (Tudò, APC B10310), for permission to carry out the study. Thanks too to Xavier Larruy, Albert Burgas, Josep Anton Ferreres Oncins, Ana Domínguez, Cristina Extremera, Irene Jiménez, Nerea Sánchez, Berta Rodríguez, Dolors Vinyoles and Marta Rodríguez for assistance in field work. We are also most grateful to Jean Marie Boutin and Denis Roux (Office National de la Chasse et de la Faune Sauvage), together with Pascal Fosty (Fédération Départementalle de Chasseurs de l’Ariège), Maxime Gaubert and Marc Druilhe (Fédération Départementalle de Chasseurs de l’Aveyron) and Pep Planes (Federació Catalana de Caça) for providing hunting data. Finally, we wish to acknowledge EURING databank (EDB) for supplying the ring data. References BirdLife International, 2004. Birds in the European Union: a status assessment. BirdLife International, Wageningen. Burfield, I., 2004. Birds in Europe. Population estimates, trends and conservation status. BirdLife Conservation Series No. 12, BirdLife International, Cambridge. Davis, P., Erard, C., Preuss, M., Tekke, N. & Tricot, J., 1966, Invasion des cailles en Europe durant l’ année 1964. Aves, 3: 65–97. Fontoura, A. P., Gonçalves, D., Guyomarc’h, J. C. & Saint–Jalme, M., 2000. La sexualité précoce des populations hivernantes de cailles des blés (Coturnix c. coturnix) au Portugal. Cahiers d’Ethologie, 20: 21–34. Gallego, S., Puigcerver, M., Rodríguez–Teijeiro, J. D.,

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Rodrigo–Rueda, F. J. & Roldán G., 1993. Algunos aspectos fenológicos y de la biología de la reproducción de la codorniz (Coturnix c. coturnix) en Cataluña (España). Historia Animalium, 2: 125–136. Gallego, S., Puigcerver, M. & Rodríguez–Teijeiro, J. D., 1997. Quail Coturnix coturnix. In: The EBCC atlas of European breeding birds: their distribution and abundance: 214–215 (W. J. M. Hagemeijer & M. J. Blair, Eds.). T. & A. D. Poyser, London. Guyomarc’h, J. C. & Belhamra, M., 1998. Les effets de la sélection sur l’expression des tendances sexuelles et migratoires chez la caille des blés (Coturnix c. coturnix). Cahiers d’Ethologie, 18: 1–16. Guyomarc’h, J. C., Guyomarc’h, C. & Saint–Jalme, M., 1989. Analyse démographique des populations de cailles des blés en Castille. Bull. Mens. Off. Natl. Chasse, 138: 34–36. Guyomarc’h, C., Lumineau, S., Vivien–Roels, B., Richard, J. & Deregnaucourt, S., 2001. Effect of melatonin supplementation on the sexual development in European quail (Coturnix coturnix). Behavioural Processes, 53(1–2): 121–130. Guyomarc’h, J. C., 2003. Elements for a Common Quail (Coturnix c. coturnix) management plan. Game and Wildlife Science, 20: 1–92. Heim de Balsac, H. & Mayaud, N., 1962. Les oiseaux du Nord–Ouest de l’Afrique. P. Chevalier Ed., Paris. Macchio, S., Messineo, A., Licheni, D. & Spina, F., 1999. Atlante della distribuzione geografica e stagionale degli uccelli inanellati in Italia negli anni 1980–1994. Istituto Nazionale per la Fauna Selvatica, Bologna. Munteanu, D. & Maties, M., 1974. The seasonal movements of the quail in Romania. Travaux du Museum d’Histoire Naturelle 'Grigore Antipa', 15: 365–380. Perennou, C., 2009. European Union Management Plan 2009–2011. Common quail, Coturnix coturnix. Technical Report, 2009–032. European Commission, Brussels. Puigcerver, M., Rodríguez–Teijeiro, J. D. & Gallego, S., 1989. ¿Migración y/o nomadismo en la codorniz (Coturnix c. coturnix)? Etología, 1: 39–45. Rodríguez–Teijeiro, J. D., Puigcerver, M., Rodrigo– Rueda, F. J. & Gallego, S., 1996., La codorniz (Coturnix coturnix) y la media veda en España. Revista Florestal, 9(1): 137–148. Rodríguez–Teijeiro, J. D., Puigcerver, M. & Gallego, S., 2004. Guatlla Coturnix coturnix. In: Atles dels ocells nidificants de Catalunya 1999–2002: 112–113 (J. Estrada, V. Pedrocchi, L. Brotons, & S. Herrando, Eds.). Institut Català d’Ornitologia (ICO)/Lynx Edicions, Barcelona. Rodríguez–Teijeiro, J. D., Barroso, A., Gallego, S., Puigcerver, M. & Vinyoles, D., 2006. Orientation–cage experiments with the European Quail during the breeding season and autumn migration. Canadian Journal of Zoology, 84: 887–894. Rodríguez–Teijeiro, J. D., Sardà–Palomera, F., Nadal, J., Ferrer, X., Ponz, C. & Puigcerver, M., 2009. The effects of mowing and agricultural landscape management on population movements of the common quail. Journal of Biogeography, 36: 1891–1898. Saint Jalme, M. & Guyomarc´h, J. C., 1995. Plum-

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age development and moult in the European Quail Coturnix c. coturnix: criteria for age determination. Ibis, 137: 570–581. Seiwert, C. M. & Adkins–Regan, E., 1998. The foam production system of the male japanese quail: characterization of structure and function. Brain, Behaviour and Evolution, 52(2): 61–80. Sinclair, A. R. E., 1984. The function of distance movements in vertebrates. The ecology of animal movement. Clarendon Press, Oxford. Weatherhead, P. J. & Greenwood, H., 1981. Age and condition bias of decoy–trapped birds. Journal of

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Field Ornithology, 52: 10–15. Wernham, C. V., Toms, M. P., Marchant, J. H., Clark, J. A. Siriwardena, G. M. & Baillie, S. R., (Eds.) 2002. The Migration Atlas: movements of the birds of Britain and Ireland. T. & A. D. Poyser, London. Zduniak, P. & Yosef, R., 2008. Age and sex determine the phenology and biometrics of migratory Common Quail (Coturnix coturnix) at Eilat, Israel. Ornis Fennica, 85: 37–45. Zuckerbrot, Y. D., Safriel, U. N. & Paz, U., 1980. Autumn migration of quail Coturnix coturnix at the north coast of the Sinai Península. Ibis, 122: 1–14.

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Determining population trends and conservation status of the common quail (Coturnix coturnix) in Western Europe M. Puigcerver, F. Sardà–Palomera & J. D. Rodríguez–Teijeiro

Puigcerver, M., Sardà–Palomera, F. & Rodríguez–Teijeiro, J. D., 2012. Determining population trends and conservation status of the common quail (Coturnix coturnix) in Western Europe. Animal Biodiversity and Conservation, 35.2: 343–352. Abstract Determining population trends and conservation status of the common quail (Coturnix coturnix) in Western Europe.— In this paper we review the conservation status and population trends of the common quail (Coturnix coturnix) from 1900 to the present. Data are sometimes contradictory with regard to the status of this species as it has some features that make it difficult to produce reliable population estimates. Recent data clearly suggest, either at a local scale or at a trans–national scale, that the Atlantic common quail populations have remained stable in the last two decades, and that restocking practices with farm–reared quails (hybrids with the Japanese quail, Coturnix japonica) do not affect our estimates. The complex movement patterns showed by this species require special attention. Analysis of ring recoveries can give important information, especially about the nomadic movement of quails in search of suitable habitats after the destruction of winter cereal crops due to harvesting. Thus, when developing a breeding distribution model for this species, continuously updated information on seasonal habitat and weather must be included for optimal prediction. Including fortnightly data of vegetation indices in distribution models, for example, has shown good results. Obtaining reliable predictions about changes in species distribution and movements during the breeding period could provide useful knowledge about the conservation status and population trends and would help in the design of future management measures. Key words: Conservation status, Population trends, Hybrids, Nomadic movements, Management. Resumen Determinación de las tendencias poblacionales y el estado de conservación de la codorniz común (Coturnix coturnix) en Europa Occidental.— En el presente estudio hacemos una revisión del estado de conservación y las tendencias poblacionales de la codorniz común (Coturnix coturnix) desde 1900 hasta nuestros días. Algunos de los datos de los que disponemos son contradictorios con respecto al estado de la especie, que presenta ciertas características que dificultan el poder proporcionar estimas poblacionales fiables. Datos recientes sugieren claramente, tanto a escala local como a escala transnacional, que las poblaciones atlánticas de codorniz común han permanecido estables en las dos últimas décadas y que la práctica de liberar codornices criadas en granjas (híbridas con la codorniz japonesa, Coturnix japonica) con finalidades cinegéticas, no afectan significativamente a nuestras estimas. Por otra parte, los complejos patrones de desplazamiento de esta especie requieren especial atención. En este sentido, el análisis de recuperaciones de anillas puede aportar información relevante, especialmente de los movimientos nomádicos de codornices a la búsqueda de hábitats adecuados, tras la destrucción de los cultivos invernales de cereales debido a la siega. Así, al desarrollar un modelo de distribución de cría para esta especie, se debe incorporar continuamente información actualizada de los cambios estacionales de hábitat y clima, con el fin de obtener unas predicciones óptimas. En este sentido, por ejemplo, la inclusión de datos quincenales de índices de vegetación en los modelos de distribución ha dado muy buenos resultados. La obtención de predicciones fiables de los cambios de la distribución de la especie y de sus desplazamientos durante la estación de cría puede ser muy útil para un mejor conocimiento del estado de conservación y las tendencias poblacionales de la especie, así como para el diseño de futuras medidas de gestión. ISSN: 1578–665X

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Palabras clave: Estado de conservación, Tendencias poblacionales, Híbridos, Desplazamientos nomádicos, Gestión. Received: 8 II 12; Conditional acceptance: 23 IV 12; Final acceptance: 16 V 12 Puigcerver, Dept. de Didàctica de les Ciències Experimentals i la Matemàtica, Fac. de Formació del Professorat, Univ. de Barcelona, Psg. Vall d’Hebron 171, 08035 Barcelona, Espanya (Spain).– F. Sardà–Palomera, Centre Tecnològic Forestal de Catalunya, Àrea de Biodiversitat, carretera Sant Llorenç de Morunys km 2 (ctra vella), 25280 Solsona, Espanya (Spain).– J. D. Rodríguez–Teijeiro, Dept. de Biologia Animal, Fac. de Biologia, Univ. de Barcelona, Avda. Diagonal 645, 08028 Barcelona, Espanya (Spain).

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Historical review of the conservation status of the common quail It is generally accepted by the scientific community that to develop suitable management and conservation policies it is necessary to have an exhaustive knowledge of the life history of the species we seek to preserve and to provide reliable population estimates (IUCN, 2001; Perennou, 2009). However, such estimates are not yet clear for the common quail, a relatively protected (Bern Convention of 1979) huntable species (Birds Directive 2009/147/EC) with an unfavourable conservation state (Bonn Convention of 1983). If we analyse the changes in conservation status over time, we can define three different stages: (1) The period from 1900 to the 1970s. Not surprisingly, during this period the status of the species remains unclear, mainly because of a lack of reliable data. However, it is accepted that a marked decrease occurred in the first half of the century in the Atlantic populations (Moreau, 1951). This is supported by a long series of data (1900–1959) in Luxembourg which suggests a dramatic decrease in quail abundance, starting at the beginning of the 20th century and finishing in the 1930s; however, a partial recovery was observed from 1947 to 1950. Later, another decreasing trend was observed (Davies et al., 1966). Moreover, in France, data collected by one hunter from 1965 to 1988 in the Lauragais region showed a decrease, especially from 1970 to 1977 (Combreau, 1992). Saint–Jalme & Guyomarc’h (1989) suggested there was a progressive reduction of wintering individuals in the Sahel after 1950, coinciding with the development of irrigated perimeters in the Maghreb, thus entailing a reduction of long–migrant phenotypes which would arrive in Europe. However, there is some evidence indicating that perhaps this decreasing trend was not so clear; in the 1940s, an increase in the number of quails was observed in some European countries, such as Germany, United Kingdom and the Scandinavian countries. Moreover, the spring passage of quails at Cape Bon (Tunisia) from 1953 to 1966 did not show a decrease (Derégnaucourt, 2000). (2) The period from 1970 to 1990. According to Birdlife International (see Burfield, 2004), in this period quail populations declined in many countries of central and northern Europe, leading to an unfavourable conservation status; populations were thus vulnerable and in large decline (Tucker & Heath, 1994). This decrease is statistically significant (Sanderson et al., 2006), but as suggested by Perennou (2009), the quantitative amplitude of the decline is unknown due to the lack of reliable pan–European estimates or indices. (3) The period from 1990 to 2000. Birdlife International (see Burfield, 2004) suggests that common quail populations were depleted due to a large historical decrease, leading to an unfavourable population status in Europe. When compared to the previous period (1970–1990), some increases were observed in northern and central Europe, whereas some declines were observed in south–eastern Europe. However, a global significant decrease was not observed (Sanderson et al., 2006). Indeed, when we restrict the distribution area to the European Union of the 25 state members,

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population trends remain stable, the threat status of the species is secure, and its conservation status is favourable (Birdlife International, 2004). These data are surprising bearing in mind that most long–distance migrants are currently declining alarmingly (Birdlife International, 2004) and farmland birds are in steep decline (Gregory et al., 2005). The common quail has the double category of being both a migratory species and a farmland bird, so it constitutes an exception to these general trends. Over the last decades, other data fail to clarify the situation because of further contradictory results (Puigcerver et al., 2004). For example, the French 'Office National de la Chasse et de la Faune Sauvage' carried out national hunting surveys in 1983 (Ferrand, 1986) and in 1988. According to the data obtained, a decrease of 50% was observed when comparing these two sampling years. In Spain, a monitoring programme of the most common bird species ('SACRE' Programme) showed a decrease in the 1998–2006 period, with an annual evolution of –6.8% (Carrascal & Palomino, 2008); from 1998 to 2010, the percentage of change reached –38% (Escandell et al., 2011). In agreement with this information, when comparing the data of the first Catalan breeding birds atlas, encompassing the period 1975–1983 (Muntaner et al., 1983) with data from the second atlas covering the period 1999–2002 (Estrada et al., 2004), a decrease of 27% in species distribution is found. On the other hand, no significanct trend was observed for quails hunted in Spain during the period 1976–2008 (Rodríguez–Teijeiro et al., 2009; Yearbook of Agro–alimentary Statistics of the Spanish Ministry of Agriculture, Fishing and Food). Furthermore, at a local scale, a long unpublished data series of common quails hunted under a constant effort in Torres de Alcanadre (Huesca Province, Spain) during 1992–2010 showed a significant increasing trend. It is also of note that no significant trends were observed in a study carried out in two Catalan breeding sites throughout the breeding seasons from 1988 to 2011 (see 'Advances towards more reliable estimates. Common quail monitoring' section). The reliability of common quail population estimates These contradictory data show that it is not easy to provide reliable estimates of common quail populations. Gregory et al. (2005) do not provide population indexes of the common quail in their study concerning the trends of farmland birds in Europe, stating that it is 'highly volatile in numbers and has an erratic migrant breeding population'. In spite of the effort carried out by Guyomarc’h (1992) to describe and to understand the structure, working and micro–evolution of the common quail populations of the Western Palearctic, it is not easy to provide reliable estimates of this species for several reasons: (1) It is (or it was) usual practice in European countries where the common quail is a popular game species to restock with farm–reared quails, which were often hybrids of common quail (Coturnix coturnix) and Japanese

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11ºW 10ºW 9ºW 8ºW 7ºW 6ºW 5ºW 4ºW 3ºW 2ºW 1ºW 0º 1ºE 2ºE 3ºE 4ºE 5ºE

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3ºW–0º

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a = 76.5º r = 0.947 P = 0.004 n = 5

a = 94º r = 0.302 P < 0.01 n = 75

a = 153º r = 0.1 P = 0.655 n = 43

a = 249º r = 0.477 P = 0.024 n = 16

a = 258º r = 0.871 P < 0.001 n = 49

a = 258º r = 0.871 P < 0.001 n = 13

Fig. 1. Recoveries of Spanish ringed individuals (white dots in the map and triangles in the circular diagrams) and preferred directions, grouped in sectors of geographical longitudes within an axis east (3º)–west (–9º). Rayleigh test statistics are shown: n. Sample size; α. Mean angle of the distribution; r. Rayleigh test statistic; and p. Significant value of the test. (Black dots are ring recoveries of quails ringed in other European countries from Rodríguez–Teijeiro et al., 2009.) Fig. 1. Recuperación de individuos hispánicos anillados (con puntos en el mapa y triángulos en los diagramas circulares) y direcciones preferidas, agrupadas en sectores de longitudes geográficas dentro de un eje este (3º)–oeste (–9º). Se presentan los parámetros del test de Rayleigh: n. Tamaño de la muestra; α. Ángulo medio de distribución; r. Estadístico del test de Rayleigh; p. Valor significativo del test. (Los puntos negros representan las recuperaciones de codornices anilladas en otros países europeos de Rodríguez–Teijeiro et al., 2009.)

quail (Coturnix japonica) (Amaral et al., 2007; Barilani et al., 2005; Chazara et al., 2006; Sanchez–Donoso et al., 2012). These practices are currently forbidden in most European countries (2002 in France and Portugal, 2007 in Spain, for example) but they might constitute a distortion factor when trying to census native common quail populations due to the large number of restocked individuals and to the difficulties of distinguishing common quails and farm–reared hybrids from their phenotypes; thus, in Catalonia (NE Spain), male common quail populations have been estimated between 5,374 and 20,847 individuals (Estrada et al., 2004), and during the period 1990–2006, a total of 1,161,113 farm–reared

quails were restocked (65,295 individuals restocked per year, see Puigcerver et al., 2007). (2) The common quail is extremely mobile. It is not only a partial migrant species, with long–migrant and short–migrant phenotypes, together with sedentary ones (Belhamra, 1997; Guyomarc’h & Belhamra, 1998; Saint–Jalme 1990; Saint–Jalme et al., 1988), but also a nomadic species, in search of suitable but ephemeral habitats which mainly comprise cereal crops, such as wheat and barley. The ripening of these cereal crops varies in latitude and in altitude, so habitats placed at lower latitudes and altitudes are destroyed (due to harvesting) earlier than those located in higher latitudes and

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altitudes. The result is that quails match their biological cycle to those of cereal crops, and their movements fit the temporal and spatial variations which constantly and predictably affect the cereal crops (Puigcerver et al., 1989; Rodríguez–Teijeiro et al., 2009). Thus, the analyse of recoveries of common quail individuals ringed in Spain showed that the timing of harvesting, combined with geographical relief, can act as a funnel forcing quail populations to concentrate with pre–migratory dispersive movements in certain areas (Rodríguez–Teijeiro et al., 2009; fig. 1; see 'ringing recoveries section'). There is another type of movement, carried out by unpaired males in search of females, which has been described as the 'Don Juan movement' (Rodríguez–Teijeiro et al., 2006); generally, these are movements of 50 km or more (own unpublished data). These three movements cause a constant inflow and outflow of males in the breeding areas throughout the breeding season, resulting in a turnover ratio of almost 95% in less than 15 days (Rodríguez–Teijeiro et al., 1992). As a consequence, it is usual to capture four times more individuals than those daily censused during the breeding season (Rodríguez–Teijeiro et al., 1992) despite using a capture method which has an effectiveness of 50% (Gallego et al., 1993). (3) Last but not least, we should not forget that the breeding cycle of the common quail occurs inside dense cereal crops and goes unnoticed; unpaired males can be acoustically detected, but females and breeding pairs remain visually and acoustically invisible to the observer (Guyomarc’h, 2003). Advances towards more reliable estimates. Common quail monitoring To address this complicated situation, a specific census methodology has recently been proposed for the common quail (Rodríguez–Teijeiro et al., 2010). This method entails continuous monitoring from year to year based on the census and capture of calling males throughout the breeding season (once a week in ten count points, following Bibby et al., 2000), which provides reliable information on density and phenology. This method is more efficient (more individuals are detected) than that provided by Guyomarc’h et al. (1998), which does not involve capture. As the harvesting period occurs when broods remain with the female in most breeding areas, monitoring carried out during harvest is also a useful methodology. It allows the number of females that have bred in the sampled areas to be censused, and the size and approximate age of broods to be known. Monitoring hunting bags may also provide useful and complementary information on the sex and age proportions of common quail populations during hunting periods. This methodology was applied at a local scale in two breeding sites in Catalonia, Northeast Spain (Figuerola del Camp and Alp), from 1988 to 2011, and at a trans–national scale in 10 different breeding sites in four countries (Morocco, Portugal, Spain and France) considered central to the Atlantic population (Gallego et al., 1997), from 2005 to 2009 (fig. 2).

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Montbel

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La Cavalerie Cabañeros– Valdesogo Mirandela

Alp Figuerola del Camp

Maranhao Sanlúcar la Mayor– Aznalcollar

Fki–Ben–Salah

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250

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Fig. 2. Breeding sites in Morocco, Portugal, France and Spain where the common quail was monitored using the method of Rodríguez–Teijeiro et al. (2010). Fig. 2. Lugares de cría en Marruecos, Portugal, Francia y España, en los que la codorniz común se monitorizó mediante el método de Rodríguez– Teijeiro et al. (2010).

With regard to the census in Catalonia, the abundance index (annual average number of quails censused per day sampled at weekly intervals throughout the breeding season) clearly shows that common quail populations have marked interannual fluctuations, but there was no significant trend in either of the two populations sampled over the last 24 years (fig. 3). At a trans–national scale, when globally analysing the 10 breeding sites of the four countries of the Atlantic population from 2005 to 2009, no significant trends were observed in the average modal number of male quails detected during the breeding seasons (as a surrogate of an abundance index), either in the breeding sites or in the countries analysed (Rodríguez– Teijeiro et al., 2010; fig. 4). This modal value ranged from two individuals in 2005 in Figuerola del Camp (Spain) to 130 in 2006 in Fki–Ben–Salah (Morocco). Thus, the results from local data, supported by the results from the trans–national study, suggest that the

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Modal number of male quails

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90 80 70 60 50 40 30 20 10 0 2004

2005

2006

2007 Year

2008

2009

2010

Fig. 3. Annual average of the modal number of male quails (± SE) censused in the breeding seasons of 2005–2009 at the breeding sites reported in fig. 2. Fig. 3. Promedio anual del número modal de codornices macho (± EE) censadas en las estaciones de cría de 2005–2009 en los lugares de cría incluidos en la fig. 2.

Atlantic common quail populations remain stable, as suggested by Fontoura & Gonçalves (1998) and Burfield (2004). This is contrary to the data presented by Ferrand (1986) in conjunction with the data of the French national hunting survey of 1998, and to findings by Muntaner et al. (1983), Estrada et al. (2004), Carrascal & Palomino (2008) and Escandell et al. (2011). Detection of hybrids As the turnover ratio of males is around 95% in less than 15 days (Rodríguez–Teijeiro et al., 1992) in northeast Spain, it is necessary not only to census individuals, but also to capture and ring them. A common method to capture quails during the breeding season consists in attracting males towards a net horizontally extended over a cereal crop, with the aid of an electronic female decoy, forcing individuals approaching it to fly and thus trapping them in the net. This capture method ensures an effectiveness of 50% (Gallego et al., 1993); its application allows us to detect hybrid males by the differences in their call structure (Collins & Goldsmith, 1998), as learning has no influence on the development of vocalisations (Konishni & Nottebom, 1969; Baptista, 1996), which are very stereotyped in the male common quail (Schleidt & Shalter, 1973). However, this method underestimates the proportion of hybrids by 50% on the basis of nuclear DNA analyses (Puigcerver et al., 2007). The first hybrid detected in Catalonia (Northeast Spain) was in 1990 (Rodríguez–Teijeiro et al., 1993), coinciding with the beginning of restocking practices involving farm–reared hybrids. Over the next 21 years, the number of censused hybrids (on the basis of call structure) was, on average, 2.2% of the total num-

ber of captured individuals, yielding a more reliable percentage estimate of 4.4 ± 0.66% hybrids, bearing in mind the underestimation inherent in the call method. During this period (1990–2011), there was no significant increasing trend in % hybrids (regression coefficient = 0.18 ± 0.17; R2 = 0.001; F1,20 = 0.028; p > 0.05; fig. 4), suggesting that hybrids are not a relevant distortion factor in the number of censused individuals (Puigcerver et al., 2007). This result could 'a priori' be unexpected, because thousands of hybrid quails were sold each year until the late 1990s by professional game breeders in Spain, France and Italy for restocking prior to the opening of the hunting season (Guyomarc’h, 2003; Puigcerver et al., 2004). Under these conditions, a rapid increase in the proportion of hybrids in common quail populations would be expected, because once hybridization has begun (as Derégnaucourt et al., 2002 suggest), it is difficult to stop, especially if hybrids are fertile and mate both with other hybrids and with parental individuals (Allendorf et al., 2001). After a few generations, this process would result in a hybrid swarm in which essentially all individuals are of hybrid origin (Huxel, 1999; Allendorf et al., 2001), leading to a collapse of the pure migratory common quail population. The low proportion of Japanese quails or hybrids captured (less than 5%) found in the last 21 years in Catalonia clearly suggests an extremely high mortality rate of released. It has been suggested that these birds are badly adapted to the wild, lacking the ability to defend themselves against cold, to forage for and select food, and to display anti–predator behaviour. (Guyomarc’h, 2003). This hypothesis is supported by data found in farm–reared red–legged partridges, where the global survival rate of 20 radio–tagged individuals was 15% three months after the release (Duarte &

Censused quails per sampling days

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50 45 40 35 30 25 20 15 10 5 0 1985

1990

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Figuerola

2000 Year Alp

2005

2010

2015

Average

Fig. 4. Male quails in two breeding sites of Catalonia: Figuerola del Camp (41.23 N, 1.17 E) and Alp (42.23 N, 1.53 E) during the period 1998–2011, censused in Catalonia using the method of Rodríguez–Teijeiro et al. (2010). Grey line: average of censused males across sites. Fig. 4. Codornices macho en dos lugares de cría de Cataluña: Figuerola del Camp (41,23 N, 1,17 E) y Alp (42,23 N; 1,53 E) durante el periodo 1998–2011 censadas en Cataluña utilizando el método de Rodríguez– Teijeiro et al. (2010). Línea gris: promedio de los machos censados en los emplazamientos.

Vargas, 2004). Furthermore, hunting practices may be a significant mortality factor for these restocked individuals; Guyomarc’h (2003) reports that, in a 64,000 ha sampling area of Haute–Garonne, 4,950 quails were hunted and 75% of them were restocked individuals. In spite of these past extensive restocking practices, our results show that they do not constitute a relevant distortion factor when censusing native breeding populations of common quail. Ringing recoveries To understand the maintenance of large–scale movement patterns, it can be helpful to explore the spatio–temporal dynamics of the resources making up the niche of a species. In particular, it is important to understand how predictability, variability and other statistical properties vary across space and time (Jonzén et al., 2011). In the case of the common quail, two of the described movements (altitudinal and aestival movements) are associated with habitat seasonality. Thus, cereal crops are, in general, ephemeral but predictable habitats; only meteorological variables provide a certain degree of unpredictability with regard to their capacity for holding common quail populations. This capacity is extended in time proportionally to the altitude of the site, so associations between the movements of the species throughout the breeding season and specific patterns of altitudinal landscapes can be predicted.

With regard to recoveries of individuals ringed in Spain (EURING database), results have clearly shown that in the northern half Spain, individuals ringed in the west have a preferred direction towards east, whereas individuals ringed in the east show a preferred direction towards the west. However, those individuals ringed between longitudes 5º W to 3º W do not show any preferred direction (fig. 1). As the harvesting date for barley in Spain is positively correlated with elevation and latitude, harvesting in combination with some geographical relief acts as a funnel, forcing quail populations to concentrate with pre–migratory dispersive movements on the Castilian Plateau, particularly in the province of Burgos (Rodríguez–Teijeiro et al., 2009). Thus, the analysis of ring recoveries is a useful method to understand movement patterns of common quails during the breeding and pre–migratory season. Conclusions and management recommendations It is clear that it is very difficult to determine population status in the case of the common quail. Generation of reliable population estimates for bird species is an important step towards determining their conservation status so as to develop appropriate conservation policies (IUCN, 2001), and as population monitoring is a top priority action recommended by the European Union

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Management Plan for the species (Guyomarc’h, 2003; Perennou, 2009), it would be desirable to generalise the census methodology proposed in Rodríguez–Teijeiro et al. (2010) to the countries of the quail´s distribution area. This would strengthen our findings with respect to the stability of the Atlantic population, which is not the result of restocking practices. The use of ring recovery data can reveal valuable information about common quail population movements. In the case of Spain, these data showed the existence of pre–migratory movements orientated towards the Castilian Plateau, where Spanish common quail populations concentrate before the opening of the hunting period. This information can be used to identify priority areas that deserve special attention in terms of conservation. Finally, the development of a species distribution model is a promising tool for the management of the common quail. This model should be based on specific monitoring, reproducing the species dynamics, and taking into account temporal replicates per site. This would allow the inclusion of a larger number of conditions to estimate environment —species relationships, as suggested in Sardà–Palomera et al. (2012). Data from meteorological stations close to monitoring locations on temperature and precipitation during the breeding season would be desirable to relate male occurrence and densities with variations of climate and weather conditions. As the life cycle of this species is closely linked to the herbaceous farmland habitat, information on changes in the temporal development of vegetation should be included in the model to reflect variations in habitat suitability. One means of capturing an element of the phenology of these dynamic landscapes is through the use of vegetation indices (Pettorelli et al., 2005). These indices provide information about photosynthetic activity, vegetation cover and structure, and they are continuously collected via satellite via remote sensors. Such a model may indicate the suitability of habitats for quails in space and time, and may also help to predict possible conflicts arising from agricultural practices and from the start date of the hunting season. Taken together, the data presented here may be useful for designing management and conservation measures for this species in order to improve its current and future conservation status. Acknowledgements The authors are most grateful to the Catalan 'Direcció General de la Recerca' (2009–SGR–481), the Spanish Science Ministry (CGL2004–05308/BOS and CGL2007–63199 projects), the 'Real Federación Española de Caza' (2005, 2006), the 'Fundación Biodiversidad' (Environmental Spanish Ministry, 2005 and 2006), the Federación de Caza de Euskadi (2007 to 2009) and to the Catalan Department of Environment, which provided financial support. EURING provided data of common quail recoveries. The family Ferreres kindly gave us a long data series of hunted quails in Torres de Alcanadre (1992–2010) under a constant hunting effort. Inés Sánchez–Donoso, Ana

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Domínguez, Cristina Extremera, Irene Jiménez, Nerea Sánchez, Dolors Vinyoles, Marisa García and Marta Rodríguez helped us in field work. Finally, we acknowledge EURING databank (EDB) which supplied the ring data. References Allendorf, F. W., Leary, R. F., Spruell, P. & Wengburg, J. K., 2001. The problems with hybrids: setting conservation guidelines. Trends in Ecology and Evolution, 16: 613–622. Amaral, A. J., Silva, A. B., Grosso, A. R., Chikhi, L., Bastos–Silveira, C. & Dias, D., 2007. Detection of hybridization and species identification in domesticated and wild quails using genetic markers. Folia Zoologica, 56(3): 285–300. Baptista, L. F., 1996. Nature and its nurturing in avian vocal development. In: Ecology and Evolution of Acoustic Communication in Birds: 39–60 (D. E. Kroodsma & E. H. Miller, Eds.). Cornell Univ. Press, Ithaca. Barilani, M., Deregnaucourt, S., Gallego, S., Galli, L., Mucci, N., Piombo, R., Puigcerver, M., Rimondi, S., Rodríguez–Teijeiro, J. D., Spanò, S. & Randi, E., 2005. Detecting hybridization in wild (Coturnix c. coturnix) and domesticated (Coturnix c. japonica) quail populations. Biological Conservation, 126: 445–455. Belhamra, M., 1997. Les effets de la sélection sur la variabilité des tendances sexuelles et migratoires dans une population captive de Caille des blés (Coturnix coturnix). Contribution à la connaissance des processus micro–évolutifs dans les populations naturelles. Ph. D. Thesis, Univ. of Rennes. Guyomarc’h, J. C. & Belhamra, M., 1998. Les effets de la sélection sur l’expression des tendances sexuelles et migratoires chez la caille des blés (Coturnix c. coturnix). Cahiers Ethol., 18: 1–16. Bibby, C. J., Burgess, N. D., Hill, D. A. & Mustoe, S. H., 2000. Bird census techniques. Academic Press, London. BirdLife International, 2004. Birds in the European Union: a status assessment. Birdlife International, Wageningen. Burfield, I., 2004. Birds in Europe. Population estimates, trends and conservation status. BirdLife Conservation Series, 12. BirdLife International, Cambridge. Carrascal, L. M. & Palomino, D., 2008. Las aves comunes reproductoras en España. Población en 2004–2006. SEO/BirdLife, Madrid. Chazara, O., Lumineau, S., Minvielle, F., Roux, D., Feve, K., Kayang, B., Boutin, J. M., Vignal, A., Coville, J. L. & Rognon, X., 2006. Étude des risques d’introgression génétique de la caille des blés (Coturnix coturnix coturnix) par la caille japonaise (Coturnix coturnix japonica): comparaison et intégration des données comportementales et moléculaires obtenues dans le sud–est de la France. Les actes du BRG, 6: 317–334. Collins, S. A. & Goldsmith, A. R., 1998. Individual and species differences in quail calls (Coturnix c. japonica, Coturnix c. coturnix and a hybrid). Ethology, 104: 977–990. Combreau, O., 1992. Etudes des variations saison-

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3.1. IUCN Species Survival Commission, Gland. Jonzén, N., Knudsen, E., Holt, R. D. & Saether, B. E., 2011. Uncertainty and predictability: the niches of migrants and nomads. In: Animal Migration: 91–109 (E. J. Milner–Ulland, J. M. Fryxell & A. R. E. Sinclair, Eds.). Oxford Univ. Press, Oxford. Konishi, M. & Nottebohm, F., 1969. Experimental studies in the ontogeny of avian vocalizations. In: Bird Vocalizations: 29–48 (R. A. Hinde, Ed.). Cambridge Univ. Press, Cambridge. Moreau, R. E., 1951. The British status of the quail and some problems of its biology. British Birds, 44: 259–276. Muntaner, J., Ferrer, X. & Martínez–Vilalta, A., 1983. Atlas dels ocells nidificants de Catalunya i Andorra. Ketres Ed., Barcelona. Perennou, C., 2009. European Union Management Plan 2009–2011. Common quail, Coturnix coturnix. Technical Report, 2009–032. European Commission, Brussels. Pettorelli, N., Vik, J. O., Mysterud, A., Gaillard, J. M., Tucker, C. J. & Stenseth, N. C., 2005. Using the satellite–derived NDVI to assess ecological responses to environmental change. Trends in Ecology and Evolution, 20: 503–510. Puigcerver, M., Rodríguez–Teijeiro, J. D. & Gallego, S., 1989. ¿Migración y/o nomadismo en la codorniz (Coturnix c. coturnix)? Etología, 1: 39–45. – 2004. Codorniz común. In: Libro rojo de las aves de España: 189–193 (A. Madroño, C. González & J. C. Atienza, Eds.). Dirección General para la Biodiversidad and SEO/BirdLife, Madrid. Puigcerver, M., Vinyoles, D. & Rodríguez–Teijeiro, J. D., 2007. Does restocking with Japanese quail or hybrids affect native populations of common quail Coturnix coturnix? Biological Conservation, 136(4): 628–635. Rodríguez–Teijeiro, J. D., Puigcerver, M. & Gallego, S., 1992. Mating strategy in the European quail (Coturnix c. coturnix) revealed by male population density and sex–ratio in Catalonia (Spain). Gibier Faune Sauvage, 9: 377–386. Rodríguez–Teijeiro, J. D., Rodrigo–Rueda, F. J., Puigcerver, M., Gallego, S. & Nadal, J., 1993. Codornices japonesas en nuestros campos. Trofeo, 277: 48–52. Rodríguez–Teijeiro, J. D., Barroso, A., Gallego, S., Puigcerver, M. & Vinyoles, D., 2006. Orientation–cage experiments with the European Quail during the breeding season and autumn migration. Canadian Journal of Zoology, 84: 887–894. Rodríguez–Teijeiro, J. D., Sardà–Palomera, F., Nadal, J., Ferrer, X., Ponz, C. & Puigcerver, M., 2009. The effects of mowing and agricultural landscape management on population movements of the common quail. Journal of Biogeography, 36: 1891–1898. Rodríguez–Teijeiro, J. D., Sardà–Palomera, F., Alves, I., Bay, Y., Beça, A., Blanchy, B., Borgogne, B., Bourgeon, B., Colaço, P., Gleize, J., Guerreiro, A., Maghnouj, M., Rieutort, C., Roux, D. & Puigcerver, M., 2010. Monitoring and management of common quail Coturnix coturnix populations in their Atlantic distribution area. Ardeola, 57(Special): 135–144. Saint–Jalme, M., 1990. La reproduction chez la caille

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des blés (Coturnix c. coturnix); études expérimentales des cycles saisonniers et de la variabilité interindividuelle. Ph. D. Thesis, Univ. de Rennes I, Rennes. Saint–Jalme, M., Guyomarc’h, J. C. & Hémon, Y. A., 1988. Acquisitions récentes sur les strategies reproductrices de la caille des blés. Bull. Mens. Off. Nat. Chasse., 127: 33–36. Saint–Jalme, M. & Guyomarc’h, J. C., 1989. Recent changes in population dynamics of European Quail in the western part of its breeding range. Proc. of the International Union of Game Biologists Congress: 130–135. Sanchez–Donoso, I., Vilà, C., Puigcerver, M., Butkauskas, D., Caballero de la Calle, J. R., MoralesRodríguez, P. A. & Rodríguez–Teijeiro, J. D., 2012. Are farm–reared quails for game restocking really

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common quails (Coturnix coturnix)?: A genetic approach. PLoS ONE, 7(6): 1–8. Sanderson, F., Donald, P. F., Pain, D. J., Burfield, I. J. & Van Bommel, F. P. J., 2006. Long–term population declines in Afro–Palearctic migrant birds. Biological Conservation, 131(1): 93–105. Sardà–Palomera, F., Puigcerver, M., Brotons, L. & Rodríguez–Teijeiro, J. D., 2012. Modelling seasonal changes in the distribution of Common Quail Coturnix coturnix in farmland landscapes using remote sensing. Ibis, 154: 703-713. Schleidt, W. M. & Shalter, M. D., 1973. Stereotypy of a fixed action pattern during ontogeny in Coturnix coturnix coturnix. Z. Tierpsychology, 33: 35–37. Tucker, G. M. & Heath, M., 1994. Birds in Europe: their conservation status. BirdLife Conservation Series, 3. BirdLife International, Cambridge.

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The grey partridge in the UK: population status, research, policy and prospects N. J. Aebischer & J. A. Ewald

Aebischer, N. J. & Ewald, J. A., 2012. The grey partridge in the UK: population status, research, policy and prospects. Animal Biodiversity and Conservation, 35.2: 353–362. Abstract The grey partridge in the UK: population status, research, policy and prospects.— Numbers of grey partridges (Perdix perdix) have declined catastrophically over the last 50 years in the UK. By contrast, the Partridge Count Scheme of the Game & Wildlife Conservation Trust (GWCT) shows an 81% increase on participating UK sites since 2000. We explore the background and reasons for this conflicting picture. GWCT research has led to scientifically proven recommendations for improving the UK partridge environment, ranging from habitat requirements to predator density. The research has influenced UK government policy, which now includes one of the most conservation–oriented and flexible agri–environment schemes in Europe, allowing land managers to recover much of the cost of grey partridge habitat creation. Culling common predators is not covered by agri–environment schemes, so it is primarily shooting estates with private gamekeepers that have implemented the full package of management measures. The future fate of the grey partridge in the UK rests on the balance between the economics of agricultural production, agri–environment measures and shooting. Key words: Grey partridge, Perdix perdix, Causes of decline, Management for recovery, Agri–environment measures. Resumen La perdiz pardilla en el Reino Unido: estado de la población, investigación, gestión y perspectivas.— Durante los últimos cincuenta años, los efectivos de la perdiz pardilla (Perdix perdix) han descendido catastróficamente en el Reino Unido. Por el contrario, el Programa de Recuento de la Perdiz de la GWCT (Fundación para la Conservación de la Caza y la Fauna Salvaje) presenta un 81% de aumento desde el año 2000 en los lugares del Reino Unido en que interviene. En este estudio exploramos los antecedentes y las razones de estos resultados tan contradictorios. Las investigaciones de la GWCT han tenido como consecuencias recomendaciones científicamente demostradas para la mejora del medio ambiente de la perdiz en el Reino Unido, desde los requerimientos del hábitat hasta la densidad de depredadores. Dichas investigaciones han influido en la política gubernamental del Reino Unido, que ahora incluye uno de los proyectos de Europa más orientadas hacia la conservación y más flexible en cuanto a hábitat y agricultura, lo que permite a los gestores del territorio recuperar gran parte del hábitat costero de la perdiz pardilla. Actualmente, los proyectos sobre agricultura y medio ambiente no abarcan la selección de los depredadores más comunes, de manera que son principalmente los cotos de caza con guardabosques privados los que han aplicado todas las medidas de gestión. El futuro de la perdiz pardilla en el Reino Unido reside en el equilibrio entre la economía de la producción agrícola, las medidas agro–medioambientales, y la caza. Palabras clave: Perdiz pardilla, Perdix perdix, Causas de disminución, Gestión para la recuperación, Medidas agro–medioambientales. Received: 19 II 12; Conditional acceptance: 2 V 12; Final acceptance: 22 V 12 N. J. Aebischer & J. A. Ewald, Game & Wildlife Conservation Trust, Fordingbridge, Hampshire, SP6 1EF, UK. Corresponding author: N. J. Aebischer. E–mail: naebischer@gwct.org.uk ISSN: 1578–665X

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Introduction The grey partridge Perdix perdix is a traditional gamebird species across the whole of Europe. In the UK, the grey partridge was the most important driven game bird on lowland estates up to the Second World War, as evidenced by records of numbers shot (Tapper, 1992). From then onwards, the numbers shot dropped rapidly. The British Trust for Ornithology (BTO) publishes a national breeding population index starting in 1966, which shows that the fall in numbers shot was matched by a prolonged drop in the number of breeding pairs, down 86% by 2000 (Crick et al., 2004). Since 2000, the BTO population index has dropped by a further 40% by 2010 (Aebischer & Ewald, 2010 updated). In parallel, the Game and Wildlife Conservation Trust (GWCT) has been monitoring the abundance and breeding success of grey partridges across the UK through its Partridge Count Scheme (PCS) (Ewald et al., 2009). This too generates an index of the grey partridge breeding population (fig. 1). Between 1966 and 2000, this index shows a similar decline to the BTO index. However, there is a dramatic difference thereafter, because the PCS index increases by 81% between 2000 and 2010. This paper explores the reasons for this conflicting picture. It reviews the research that has been carried out to understand the causes of the grey partridge decline and to counteract them, the policy background against which the grey partridge story unfolds, and the efforts that have been put in place to restore numbers of the grey partridge in the UK. Research Concerns generated by the decline in grey partridge bags prompted a long–term research programme by GWCT starting in the late 1960s (Potts, 1986). Three main causes of decline were identified, linked to agricultural intensification: (1) destruction of nesting habitat, resulting in poor holding capacity; (2) pesticide–induced reduction in chick–food insects in crops, leading to poor chick survival; and (3) increased predation pressure on remaining habitat, leading to adult and nest losses. We present below the evidence linking these factors to the decline, the experimental work that confirmed their importance, and the research carried out to find solutions. Nesting habitat Grey partridges nest in rank grassy cover that conceals the nest from predators. Radio–tracking found that two–thirds of females hide their nest in linear boundary features such as the base of hedgerows, grassy banks or uncut field margins, the rest being mainly in autumn–sown cereals (Aebischer et al., 1994). Since the Second World War, the drive for greater agricultural efficiency has led to field enlargement through the removal of field boundaries. One consequence has been a reduction in the length of hedgerows by 40% over the last 60 years (Brown, 1992; Anon., 2009),

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with a consequent reduction in nesting cover; Potts (1980) estimated that 24% of post–war nesting cover had been lost by 1978. In addition, annual mowing or treatment with herbicides to prevent crop invasion by weeds (Boatman, 1992) further reduced nesting cover quality. Both Potts (1980) and Rands (1986) demonstrated a correlative link between the availability of suitable nesting cover and spring density of grey partridges. Restoring nesting habitat requires re–establishing areas of rank tussocky grass and other concealing vegetation, most simply as strips around field margins. Large fields can be subdivided by using non–permanent grass strips ('Beetle Banks'; Thomas et al., 1991) sown with tussock–forming grasses, which do not impede agricultural operations. Cutting management must ensure that tall dead grass is always present early in the season to provide nest cover (Aebischer, 1997). Chick–food availability Grey partridge chick survival is a key determinant of population change (Aebischer & Ewald, 2004). Its importance is demonstrated by figure 2, whereby average annual chick survival (from the PCS) explains over two–thirds of variation in the year–on–year change in the BTO population index. Accordingly, much research effort has been expended in understanding this crucial phase of the grey partridge life cycle. Grey partridge parents lead their chicks away from the nest after hatching. The chicks feed themselves and during the first two weeks their diet is made up overwhelmingly of insects (Ford et al., 1938; Potts, 1980, 1986). Nutrition experiments in 1964–65 showed that the high protein intake from insects was crucial for feather development and survival (Southwood & Cross, 2002). In the field, radio–tracking of females with broods showed that the chicks spent 97% of their time in cereal crops (Green, 1984), which must therefore have been their primary source of food. In corroboration, there was a strong relationship between chick survival and chick–food abundance in cereals at the farm scale (fig. 3). The abundance and availability of chick–food insects in cereal crops has changed dramatically with the advent of pesticides. Herbicides were first, introduced in the 1950s to combat weeds in crops, and by 1965 nearly all cereal fields were treated with them (Potts, 1980). This greatly reduced the abundance of arable weeds that acted as host plants for insects, and the abundance of chick–food insects halved as a result (table 1). Then, during the 1970s, insecticide use became widespread. Vickerman & Sunderland (1977) showed that insecticide applied in summer could reduce chick–food insects by over 90%. At the farm level, grey partridge chick survival was a third lower on areas of extensive insecticide use than on areas with little or no insecticide use (Aebischer & Potts, 1998). How then to restore insect abundance in cereals in a way compatible with modern farming? One answer was 'Conservation Headlands', whereby the outer six metres of the cereal crop ('headland') were treated selectively to encourage a weedy understorey accessible to partridge chicks while eliminating agriculturally damaging

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2.0

Count index (1961 = 1)

1.8 1.6 1.4 1.2 1.0 0.8 0.6 0.4 0.2 0.0 1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 Fig. 1. Annual index (with 95% confidence limits) of grey partridge pairs recorded on sites contributing to the GWCT’s Partridge Count Scheme 1961–2011, relative to the start year (1961 = 1). Updated from Aebischer & Ewald (2010). Fig. 1. Índice anual (con límites de confianza del 95%) de las parejas de perdiz pardilla registradas en lugares que contribuyeron al Programa de Recuento de la Perdiz del GWCT de 1961 a 2011, en relación con el año de inicio (1961 = 1). Actualizado de Aebischer & Ewald (2010)

0.5

30 20 10 0 –10 –20 –30 –40 10

r30 = 0.826 P < 0.001 20 30 40 50 % Chick survival (year t)

Fig. 2. Annual changes in the BTO national index of grey partridge abundance correlate closely with annual chick survival rates from the GWCT’s Partridge Count Scheme (Aebischer & Ewald, 2004). Fig. 2. Cambios anuales en el índice nacional BTO de abundancia de perdiz pardilla, que está estrechamente correlacionada con las tasas anuales de supervivencia de crías del Programa de Recuento de la Perdiz del GWCT (Aebischer & Ewald, 2004).

Chick survival rate

% Change (t to t +1)

0.4 0.3 0.2 0.1 0.0 0.0

0.2 0.4 0.6 0.8 1.0 % Chick–food abundance

Fig. 3. Annual grey partridge chick survival per farm is closely related to the average density of chick–food insects sampled in cereal crops on the same farm, for five farms in Sussex at the farm scale 1970–1992 (Aebischer, 1997). Fig. 3. La supervivencia anual de crías de perdiz pardilla por granja está estrechamente relacionada con la densidad promedio de insectos del alimento para crías provinente de los cultivos de cereales de la misma granja, para cinco granjas en Sussex a escala de la granja 1970–1992 (Aebischer, 1997).

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Table 1. Reduction in the abundance of grey partridge chick–food insects in cereal crops wrought by herbicide use (adapted from Potts, 1986): R. Reduction; CMPP+S. CMPP + Simazine; (1) Some MCPB; (2) Listed in Ewald & Aebischer (2000). Tabla 1. Reducción de la cantidad de insectos en el alimento para las crías de perdiz, debido al uso de herbicidas en las cosechas de grano (adaptado de Potts, 1986): R. Reducción; CMPP+S. CMPP + Simazine; (1) Algunos MCPB; (2) Citado en Ewald & Aebischer (2000). Herbicide DNOC

R 25%

Authority Johnson et al., 1955

2,4–D

36%

Ubrizy, 1968

MCPA(1)

47%

Southwood & Cross, 1969

Various

49%

Vickerman & O’Bryan, 1979

CMPP+S

58%

Vickerman, 1974

Various

55%

Sotherton et al., 1985

Various(2)

63%

Potts, 1986

Mean

48%

weeds (Sotherton, 1991). Overall, chick food was more than twice as abundant in Conservation Headlands than in conventionally treated headlands (table 2). An eight–year experiment, which divided farms in two and randomly allocated Conservation Headlands to one half and conventional management to the other, found that in all years grey partridge chick survival was higher with than without Conservation Headlands (Sotherton et al., 1993). With Conservation Headlands, in five of the eight years it exceeded the 35% level above which, from figure 2, population change would be positive i.e. the population would grow. Without them it barely reached that level in even one year (Aebischer, 1997). Hence the linkage between insect restoration and partridge recovery was established. Another method of providing chick food is by deliberately growing insect–rich brood–rearing crop mixtures (Aebischer, 1997). This option is attractive to farmers and landowners interested in shooting, as it also has benefits for other game birds. Refinements of this approach have led to growing parallel strips of cover that satisfy the year–round requirements of grey partridges, for instance by placing, a cereal mixture (brood–rearing) with first–year kale (winter cover) and quinoa (winter food), or with first–year kale and second–year kale (winter food) next to nesting cover. Predation pressure Because they nest on the ground, incubating female grey partridges and their eggs are vulnerable to

mammalian and avian predators. Traditionally, one of the roles of the private gamekeepers employed by UK shooting estates was to kill predators as part of gamebird husbandry. Since the Second World War, the number of gamekeepers involved in active predation control has fallen and the number now is merely a fifth of what is used to be (Tapper, 1992). Conversely, the BTO population indices for crows Corvus corone and magpies Pica pica have more than doubled over the last 40 years (Baillie et al., 2010), and a tripling of the national fox Vulpes vulpes bag (Aebischer et al., 2011) indicates a similar increase in fox numbers. The predation pressure on grey partridges is therefore greater now than it was in the past. The main issue is whether predation plays a role in population regulation rather than just removing a 'doomed surplus' sensu Errington (1956). A seven– year cross–over experiment conducted on two sites by GWCT showed conclusively that legal control of foxes, mustelids Mustela spp., brown rats Rattus norvegicus, crows and magpies not only increased the production of young grey partridges but also their spring pair density relative to no control (Tapper et al., 1996). Over three years, the effect resulted in a 3.5–fold difference in post–breeding numbers, and in a 2.6–fold difference in spring pair density. It is clear that high predation pressure had a considerable impact on grey partridge breeding density. At the same time, the study demonstrates that it was not necessary to remove all predators throughout the year, but that the selective removal of common predators specifically during the nesting period was enough for the partridge population to respond. This perhaps makes it more acceptable and feasible as a management tool. Policy The research described above did not take place in isolation, but against the background of major changes in both agricultural and environmental policy. To put the research in context, we review below how policy has changed in these two areas. Arable agriculture Since the UK joined the European Economic Community (EEC) in 1973, its agricultural policy has been closely linked to that of the EEC and to that of its successor, the European Union (EU). The main events affecting agricultural policy are detailed in table 3. At the EU level, the biggest changes are the implementation of production controls, which in the case of arable agriculture took the form of first voluntary then mandatory set–aside, the greening of the Common Agricultural Policy via Environmentally Sensitive Areas then agri–environment measures, the decoupling of production from agricultural subsidies, and latterly the abolition of mandatory set–aside quotas. At the UK level, these changes translated directly into national policy – introduction of voluntary then mandatory set–aside, designation of Environmen-

Animal Biodiversity and Conservation 35.2 (2012)

tally Sensitive Areas, introduction of Countryside Stewardship. But the UK government was also sensitive to the evidence of widespread declines in farmland birds (Fuller et al., 1995), and scarred by the huge cost of dealing with foot–and–mouth disease in 2001, when seven million animals were slaughtered. It therefore incorporated arable options into its agri–environment policies and commissioned a report to consider the future of farming and food (Curry, 2002). Its recommendations, combined with the opportunity offered under the EU Mid–term Review, led in 2005 to the almost total decoupling of subsidies from production with the Single Farm Payment scheme, tying subsidies instead to good agricultural practice and wildlife–friendly land management. At a stroke the economic incentive on UK farms changed completely. Defra statistics show that production subsidies dropped from £2,168 million in 2004 to £212 million in 2005, while uncoupled subsidies (Single Farm Payments) rose from £778 million to £2,819 million. Subsidies for set–aside, which had counted as production subsidies, were included with the Single Farm Payment under the new system.

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Table 2. Increase in the abundance of grey partridge chick–food insects in cereals through the use of Conservation Headlands (Sotherton et al., 1993): Cvs. Conventional sprayed; Cnh. Conservation headland; Chn. Change (%). Tabla 2. Aumento de insectos en el alimento a base de cereales para las crías de perdiz debido a la política de ''Conservation Headlands'' (Sotherton et al., 1993): Cvs. Rociado convencional; Cnh. Conservación Headlands; Chn. Cambio (%). Chick food

Densities/0.5 m2 Cvs

Cnh

Chn

Sawfly larvae & caterpillars 2.1

2.6

+24%

Leaf beetles

4.0

9.9 +148%

Plant bugs

10.4 36.0 +246%

All chick–food insects

29.9 68.6 +129%

Table 3. Chronological sequence of major policy decisions affecting agriculture at the level of the European Union (EU) and of the United Kingdom (UK) since 1962. Tabla 3. Secuencia cronológica de las principales decisiones de gestión que afectan a la agricultura de toda la Unión Europea (EU) y del Reino Unido (UK) desde 1962. EU

Policy

1962

EEC introduces Common Agricultural Policy

UK joins EEC

1973

1984

EU sets quota on dairy production

1985

EU introduces special aid for Environmentally Sensitive Areas (ESAs)

UK designates ESAs (first tranche)

1987

1988

EU introduces voluntary set–aside

1988

UK introduces voluntary set–aside

1991

UK launches Countryside Stewardship

1992

EU MacSharry reforms: mandatory set–aside, agri–environment measures

1992

UK introduces mandatory set–aside

1993

UK designates ESAs (second tranche)

1998

UK launches Arable Stewardship Pilot

1999

EU Agenda 2000: market subsidies reduced in favour of direct payments to farmers

UK adds arable options to Countryside Stewardship

2002

2003

EU Mid–term Review: subsidies decoupled from production (single–farm payment)

UK launches Environmental Stewardship (Entry Level, Higher Level)

2005

2008

EU CAP Health Check: set–aside abolished

2008

UK abolishes set–aside

2009

UK launches Campaign for the Farmed Environment

358

Aebischer & Ewald

Nesting habitat

160

Unpaid management Agri–environment Non–arable Set–aside

140

140 120

100

2009

2008

2007

2006

2009

2008

2007

0 2006

20 2005

20 2004

2003

2005

2004

2003

2002

Hectares

100

2002

Hectares

120

Brood–rearing habitat

160

Fig. 4. Breakdown of the amounts of nesting and brood–rearing habitats on the demonstration areas of the GWCT’s Grey Partridge Demonstration Project according to land–cover type, from 2002 to 2009. Fig. 4. Desglose de la cantidad de hábitats de nidificación y de cría en las áreas de prueba del Proyecto de Demostración de la perdiz pardilla de la GWCT, según el tipo de cubierta del suelo, desde 2002 a 2009.

At the same time, the UK government introduced a new Environmental Stewardship Scheme, which replaced previous agri–environment schemes while incorporating and extending their wildlife–sympathetic elements (Anon., 2005). The scheme is in two parts, the Entry Level Scheme open to all farmers, and the competitive Higher Level Scheme that offers more intensive targeted habitat management. The Environmental Stewardship Scheme comprises 36 management options related to farmland (41 including archaeological ones). Of these, six are directly due to GWCT research and a further 23 have been influenced by it. In its current form, it is one of the most conservation–minded and flexible agri–environment schemes in Europe. With regard to grey partridge, it can defray the cost of creating, for example, nesting cover in the form of grass buffer strips and Beetle Banks, and insect–rich brood–rearing habitat in the form of unharvested cereal strips or Conservation Headlands. Environment and biodiversity In the UK, the grey partridge has been red–listed since 1990 (Batten et al., 1990) and has also been classified as a Species of Unfavourable Conservation Status by the EU (Tucker & Heath, 1994). Since signing the Biodiversity Convention at Rio de Janeiro in 1992, the UK Government has committed itself to addressing biodiversity issues through its Biodiversity Action Plan (Anon., 1995), under which the grey partridge is listed as a priority species. The GWCT was nominated as lead partner for the grey partridge in 1996, i.e. given the responsibility for taking forward the objectives in the government’s species action plan: (1) halt the

decline by 2005; (2) ensure the population is above 160,000 pairs by 2020; and (3) maintain, and where possible enhance, the current range. This set the scene for the recovery programme that is described in the next section. Response Being Lead Partner brought responsibility but no money. However, the generosity of private individuals and companies enabled the GWCT to launch a major programme for partridge recovery (Aebischer, 2009). Because almost all UK land is privately owned, and sympathetic land management is central to partridge recovery, the cornerstone of the programme was to motivate farmers, landowners and shoot managers to address the causes of decline. The programme was thus primarily education–oriented, and relied on two main strands: (1) encouraging by example through a Grey Partridge Demonstration Project, and (2) developing the PCS network to monitor, inform and advise. Grey partridge demonstration project The aim of the project was to establish a demonstration where visitors might see for themselves the management techniques needed for grey partridges, observe the increase in numbers of grey partridges that results from the management, learn about the pitfalls and costs, and be motivated to follow suit (Aebischer, 2009). The project began in autumn 2001 and ended in spring 2010, on two areas of light arable farmland near Royston, Hertfordshire, some 65 km north of

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Fig. 5. Changes in grey partridge density in spring and after breeding on the demonstration and reference areas of the GWCT’s Grey Partridge Demonstration Project, from autumn 2001 to spring 2010. Management began in 2002. Fig. 5. Cambios en la densidad de perdiz pardilla en primavera, tras criar en las zonas de prueba y de referencia del Proyecto de Demostración de la perdiz pardilla de la GWC, desde el otoño de 2001 a la primavera de 2010. La gestión comenzó en el 2002.

London. One 996–ha area (six farm holdings) was the demonstration area, while a surrounding area of 1,311 ha (seven holdings) constituted a reference area for comparison. On the demonstration area, the GWCT employed a keeper to address the causes of decline in several ways. In cooperation with the farmers, he created habitat for nesting and brood–rearing, as well as providing overwinter cover. This relied heavily on set–aside and agri–environment schemes to cover the costs of management (Aebischer & Ewald, 2010). In the first five years, a third of nesting cover was created on non–rotational set–aside strips (sown to grass and not cultivated again) and another third on agri–environment land (fig. 4). Brood–rearing cover was grown as mixed cereal crops half on rotational set–aside and half on agri–environment land. The availability of rotational set– aside fell after the decoupling of production subsidies in 2005, and all set–aside disappeared after the zero quota in 2008. The shortfall was made up by entering land into the new Environmental Stewardship Scheme, although some brood–rearing cover in 2009 was unpaid. The keeper was also responsible for predator control, particularly during the partridge breeding season. He targeted foxes by night–shooting, small mustelids by tunnel–trapping, rats by poisoning, and corvids by shooting and Larsen–trapping. From September to March, he provided supplementary food in the form of wheat grain in hoppers placed along field margins and cover strips, to counteract any winter food shortage. He also counted and mapped grey partridges on the demonstration and reference areas

in spring and autumn, allocating them to one or other area depending on location at the time of counting. Initially, the density of grey partridge pairs was low, at under 3 pairs/km2 (fig. 5). It increased during the next five years, then remained around 15 pairs/km2, representing a sustainable 5–fold increase. The density of grey partridges in the autumn, after breeding, followed a similar pattern (fig. 5). It increased from 8 birds/km2 before management to around 80 birds/km2, a ten–fold increase. The ratio of spring to autumn bird numbers was 40–43% —a measure of combined overwinter survival and dispersal. It was lower than the usual value of around 50% (Potts, 1986), suggesting high emigration. Over the same period, increases on the surrounding reference area, from under 2 pairs/km2 to over 5 pairs/km2 in the absence of partridge–specific management, are consistent with high emigration from the demonstration area. Thus the data do not support the possibility that the increase observed on the demonstration area was reinforced by birds being attracted into it from the reference area. The success of the demonstration project offered convincing evidence that the combined package of habitat management, predator control and supplementary feeding was effective. Partridge count scheme (PCS) The PCS began in 1933 as a means of monitoring annual density and breeding success of the grey partridge on some 90 'partridge manors'. As part of its national Recovery Programme, the GWCT relaunched the scheme

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in 1998 under the banner 'Every one counts'. The aim was to increase participation and, beyond monitoring, to use the contact with farmers, landowners and keepers to encourage more and better management (Ewald et al., 2009). In August 2011, there were 1597 sites registered with the PCS from across the UK. Participants are asked to count the partridges on their land twice a year to enable the GWCT to monitor the number of breeding pairs and their productivity. To help contributors, the GWCT provides a guide to aging and sexing grey partridges in the spring and autumn. In addition, each contributor receives a spring and autumn newsletter, a pair density target based on landscape characteristics, and management feedback on how to achieve it. A series of fact sheets and leaflets address management issues in greater detail, covering the provision of nesting and brood–rearing habitat, methods of controlling predators, best use of agri–environmental subsidies and guidelines on shooting (all publicly available at http://www.gwct.org.uk/partridge). The management message is reinforced through a network of 16 local Partridge Groups, which hold at least one meeting a year open to all contributors within the area. The meetings offer the opportunity to talk about research, management and agri–environment options in the context of grey partridges, and often involve field visits to demonstrate successful management. Friendly competition is encouraged within each Partridge Group by awarding an annual prize for the best conservation effort. The net result is that PCS contributors are more likely than non–contributors to use agri–environment options that benefit grey partridges, notably Beetle Banks and Conservation Headlands (Ewald et al., 2010). Where we are now The grey partridge decline is of such magnitude that it is clearly a conservation priority (Eaton et al., 2009). Extensive research means that the causes of the decline are well understood, counter–measures have been investigated and solutions found that are not only compatible with modern agriculture but also adopted into and funded by UK agri–environmental schemes. In addition, CAP reform has alleviated economic pressure on managing cropped land by decoupling subsidies from production. These factors suggest that the habitat requirements of grey partridges, which involve the cropped as much as the uncropped parts of a farm, are more acceptable to land managers now than in the past. The culling of common predators is not covered by agri–environment schemes, so it is primarily on shooting estates with private gamekeepers that the full package of management measures can be implemented most cost–effectively, thanks to the alternative revenue stream that shooting offers (PACEC, 2006). How widely applicable are the solutions that have been deployed on the Grey Partridge Demonstration Project at Royston? In the Introduction, we highlighted the contrast between the PCS and the national picture given by the BTO population index, whereby since 2000 spring densities have almost doubled on land managed by PCS contributors, but nearly halved in the wider

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countryside. This success suggests a wide general applicability, but there are also broad landscape and climatic features that need to be taken into account. For instance, the GWCT’s Allerton Project at Loddington Farm in Leicestershire has undertaken habitat management, predator control and supplementary feeding for game in much the same way as at Royston from 1992 to 2001 (Stoate & Leake, 2002). Pheasants Phasianus colchicus, hares Lepus europaeus and songbirds all increased but grey partridges remained at very low density. This was probably because the landscape was wooded rather than open and the soil was heavy and wet rather than light and well–drained. A mapping exercise based on landscape features found that the optimal areas for grey partridge in the UK were primarily in the east, with suitability declining from east to west (Aebischer, 2009). Another factor that can prevent the recovery of grey partridges is intensive driven shooting of released red–legged partridges Alectoris rufa, because wild grey partridges are inadvertently shot during the drives. However, precautionary measures such as whistles to warn the guns when grey partridges are flushed over them are effective at reducing losses to a tolerable level (Watson et al., 2007). In the PCS, such precautions combined with sympathetic management result in grey partridge population growth even in the presence of high levels of red–legged partridge releasing and shooting (Aebischer & Ewald, 2010). There is thus no doubt about the effectiveness of the PCS feedback procedures and the face–to–face education carried out through the local Partridge Groups to motivate farmers and landowners into instigating grey partridge conservation measures. Despite the increases on the land they manage, however, PCS participants are too few to make an impact on the national downward trend of the grey partridge. Looking into the future, therefore, the task ahead is clear: the PCS needs to expand and motivate many more land and shoot managers. This is where the GWCT’s current efforts lie. In conclusion, land and shoot managers are key to grey partridge recovery, and education is crucial for raising awareness and encouraging them into sympathetic management. The future fate of the grey partridge rests on the balance between the economics of agricultural production, agri–environment measures and shooting. We believe that the different strands of the GWCT recovery programme form a package that, coupled with the government’s agricultural reforms, offers genuine hope for the recovery of the grey partridge in the UK. Acknowledgements We are honoured by the Scientific Committee’s invitation to present this paper as a plenary lecture of IUGB’s XXXth International Congress. We thank all the individual farmers, gamekeepers, landowners and scientists, as well as the organisations and companies, who have dedicated time, effort and money to the cause of the grey partridge.

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References Aebischer, N. J., 1997. Gamebirds: management of the grey partridge in Britain. In: Conservation and the Use of Wildlife Resources: 131–151 (M. Bolton, Ed.). Chapman & Hall, London. – 2009. The GWCT Grey Partridge Recovery Programme: a Species Action Plan in action. In: Gamebird 2006: 291–301 (S. B. Cederbaum, B. C. Faircloth, T. M. Terhune, J. J. Thompson & J. P. Carroll, Eds.). Warnell School of Forestry, Athens, USA. Aebischer, N. J., Blake, K. A. & Boatman, N. D., 1994. Field margins as habitats for game. In: Field Margins – Integrating Agriculture and Conservation: 95–104 (N. D. Boatman, Ed.). BCPC Monograph No. 58, BCPC Publications, Farnham. Aebischer, N. J., Davey, P. D. & Kingdon, N. G., 2011. National Gamebag Census: Mammal Trends to 2009. Game & Wildlife Conservation Trust, Fordingbridge (http://www.gwct.org.uk/ngcmammals). Aebischer, N. J. & Ewald, J. A., 2004. Managing the UK Grey Partridge Perdix perdix recovery: population change, reproduction, habitat and shooting. Ibis, 146 (Suppl. 2): 181–191. – 2010. Grey Partridge Perdix perdix in the UK: recovery status, set–aside and shooting. Ibis, 152: 530–542. Aebischer, N. J. & Potts, G. R., 1998. Spatial changes in Grey Partridge (Perdix perdix) distribution in relation to 25 years of changing agriculture in Sussex, U.K. Gibier Faune Sauvage, 15: 293–308. Anon., 1995. Biodiversity: The UK Steering Group Report. Volume 2: Action Plans. Her Majesty’s Stationery Office, London. – 2005. Environmental Stewardship: Look after your Land and be Rewarded. Rural Development Service, Department for Environment, Food and Rural Affairs, London. – 2009. Countryside Survey: England Results for 2007. Natural Environment Research Council, Swindon. Baillie, S. R., Marchant, J. H., Leech, D. I., Joys, A. C., Noble, D. G., Barimore, C., Downie, I. S., Grantham, M. J., Risely, K. & Robinson, R. A., 2010. Breeding Birds in the Wider Countryside: Their Conservation Status 2009. BTO Research Report No. 541. British Trust for Ornithology, Thetford. Batten, L. A., Bibby, C. J., Clement, P., Elliott, G. D. & Porter, R. F., 1990. Red Data Birds in Britain: Action for Rare, Threatened and Important Species. T. & A.D. Poyser, London. Boatman, N. D., 1992. Improvement of field margin habitat by selective control of annual weeds. Aspects of Applied Biology, 29: 431–436. Brown, A., 1992. The UK Environment. Department of the Environment, HMSO, London. Crick, H. Q. P., Marchant, J. H., Noble, D. G., Baillie, S. R., Balmer, D. E., Beaven, L. P., Coombes, R. H., Downie, I. S., Freeman, S. N., Joys, A. C., Leech, D. I., Raven, M. J., Robinson, R. A. & Thewlis, R. M., 2004. Breeding Birds in the Wider Countryside: Their Conservation Status 2003. BTO Research Report No. 353. British Trust for Ornithology, Thetford.

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Curry, D., 2002. Farming and Food: A Sustainable Future. Policy Commission on the Future of Farming and Food, Cabinet Office, London. Eaton, M. A., Brown, A. F., Noble, D. G., Musgrove, A. J., Hearn, R. D., Aebischer, N. J., Gibbons, D. W., Evans, A. D. & Gregory, R. D., 2009. Birds of Conservation Concern 3: the population status of birds in the United Kingdon, Channel Islands and Isle of Man. British Birds, 102: 296–341. Errington, P. L., 1956. Factors limiting vertebrate populations. Science, 124: 304–307. Ewald, J. A. & Aebischer, N. J., 2000. Trends in pesticide use and efficacy during 26 years of changing agriculture in southern England. Environmental Monitoring and Assessment, 64: 493–529. Ewald, J. A., Aebischer, N. J., Richardson, S. M., Grice, P. V. & Cooke, A. I. 2010. The effect of agri–environment schemes on grey partridges at the farm level in England. Agriculture Ecosystems and Environment, 138: 55–63. Ewald, J. A., Kingdon, N. G. & Santin–Janin, H., 2009. The GWCT Partridge Count Scheme: a volunteer–based monitoring and conservation promotion scheme. In: Gamebird 2006: 27–37 (S. B. Cederbaum, B. C. Faircloth, T. M. Terhune, J. J. Thompson & J. P. Carroll, Eds.). Warnell School of Forestry, Athens, USA. Ford, J., Chitty, H. & Middleton, A. D., 1938. The food of partridge chicks (Perdix perdix L.) in Great Britain. Journal of Animal Ecology, 7: 251–265. Fuller, R. J., Gregory, R. D., Gibbons, D. W., Marchant, J. H., Wilson, J. D., Baillie, S. R. & Carter, N., 1995. Population declines and range contractions among lowland farmland birds in Britain. Conservation Biology, 9: 1425–1441. Green, R. E., 1984. The feeding ecology and survival of partridge chicks (Alectoris rufa and Perdix perdix) on arable farmland in East Anglia. Journal of Applied Ecology, 21: 817–830. Johnson, C. G., Dobson, R. M., Southwood, T. R. E., Stephenson, J. W. & Taylor, L. R., 1955. Preliminary observations on the effect of weed killer DNOC on insect populations. In: Rothamsted Experimental Station Report for 1954: 129–130. Lawes Agricultural Trust, Harpenden. PACEC, 2006. The Economic and Environmental Impact of Sporting Shooting in the UK. Public and Corporate Economic Consultants, London. Potts, G. R., 1980. The effects of modern agriculture, nest predation and game management on the population ecology of partridges (Perdix perdix and Alectoris rufa). Advances in Ecological Research, 11: 1–79. – 1986. The Partridge: Pesticides, Predation and Conservation. Collins, London. Rands, M. R. W., 1986. Effect of hedgerow characteristics on partridge breeding densities. Journal of Applied Ecology, 23: 479–487. Sotherton, N. W., 1991. Conservation Headlands: a practical combination of intensive cereal farming and conservation. In: The Ecology of Temperate Cereal Fields: 373–397 (L. G. Firbank, N. Carter, J. F. Derbyshire & G. R. Potts, Eds.). Blackwell Scientific Publications, Oxford.

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Sotherton, N. W., Rands, M. R. W. & Moreby, S. J., 1985. Comparison of herbicide–treated and untreated headlands for the survival of game and wildlife. In: 1985 British Crop Protection Conference–Weeds: 991–997. British Crop Protection Council, Farnham. Sotherton, N. W., Robertson, P. A. & Dowell, S. D., 1993. Manipulating pesticide use to increase the production of wild game birds in Britain. In: Quail III: National Quail Symposium: 92–101 (K. E. Church & T. V. Dailey, Eds.). Kansas Department of Wildlife and Parks, Pratt. Southwood, T. R. E. & Cross, D. J., 1969. The ecology of the partridge III. Breeding success and the abundance of insects in natural habitats. Journal of Animal Ecology, 38: 497–509. – 2002. Food requirements of grey partridge Perdix perdix chicks. Wildlife Biology, 8: 175–183. Stoate, C. & Leake, A., 2002. Where the Birds Sing: 10 Years of Conservation on Farmland. Game Conservancy Trust & Allerton Research and Educational Trust, Fordingbridge. Tapper, S. C., 1992. Game Heritage: An Ecological Review from Shooting and Gamekeeping Records. Game Conservancy Ltd, Fordingbridge. Tapper, S. C., Potts, G. R. & Brockless, M. H., 1996. The effect of an experimental reduction in predation pressure on the breeding success and population density of grey partridges (Perdix perdix). Journal

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of Applied Ecology, 33: 965–978. Thomas, M. B., Wratten, S. D. & Sotherton, N. W., 1991. Creation of island habitats in farmland to manipulate populations of beneficial arthropods– predator densities and emigration. Journal of Applied Ecology, 28: 906–917. Tucker, G. M. & Heath, M. F., 1994. Birds in Europe: Their Conservation Status. Birdlife International (Birdlife Conservation Series no. 3), Cambridge. Ubrizy, G., 1968. Long–term experiments on the flora–changing effect of chemical weed killers in plant communities. Acta Agronomica Academiae Scientiarum Hungaricae, 17: 171–193. Vickerman, G. P., 1974. Some effects of grass weed control on the arthropod fauna of cereals. In: Proceedings of the 12th British Weed Control Conference: 929–939. British Crop Protection Council, London. Vickerman, G. P. & O’Bryan, M., 1979. Partridges and insects. In: Annual Review for 1978: 35–43. Game Conservancy, Fordingbridge. Vickerman, G. P. & Sunderland, K. D., 1977. Some effects of dimethoate on arthropods in winter wheat. Journal of Applied Ecology, 14: 767–777. Watson, M., Aebischer, N. J., Potts, G. R. & Ewald, J. A., 2007. The relative effects of raptor predation and shooting on overwinter mortality of grey partridges in the United Kingdom. Journal of Applied Ecology, 44: 972–982.

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Restoration of a wild grey partridge shoot: a major development in the Sussex study, UK J. A. Ewald, G. R. Potts & N. J. Aebischer

Ewald, J. A., Potts, G. R. & Aebischer, N. J., 2012. Restoration of a wild grey partridge shoot: a major development in the Sussex study, UK. Animal Biodiversity and Conservation, 35.2: 363–369. Abstract Restoration of a wild grey partridge shoot: a major development in the Sussex study, UK.— The scientific basis of wild grey partridge management has been known for a generation. This includes controlling nest predators, providing nesting cover, having sufficient insect food for chicks and appropriate rates of shooting. More recently, measures such as providing food for adult birds and habitats for protection from birds of prey have also been considered important. Habitat provision can be expensive, but in the UK costs can be partially recovered through governmental agri–environment schemes. The landowner still needs to pay for the essential gamekeeper. Since 2003/04, one part of the Game & Wildlife Conservation Trust’s (GWCT) Sussex Study area has put these principles of environmental management into practice with the aim of restoring a wild grey partridge shoot to this part of Southern England. Results have been impressive, with the spring pair density increasing from 0.3 pairs/100 ha in 2003 to nearly 20 pairs/100 ha in 2010 on an area of just over 10 km2. Over the past two years a wild grey partridge shoot has taken place, and the landowner and his team have gained national recognition for their conservation work. Key words: Grey partridge, Perdix perdix, Predator control, Agri–environment measures, Population recovery. Resumen Restauración de la caza de la perdiz pardilla: un importante progreso en el estudio de Sussex, Reino Unido.— Desde hace una generación se conoce la base científica de la gestión de la perdiz pardilla. Ésta incluye el control de los depredadores de nidos, la provisión de material para la nidificación, tener suficientes insectos para alimentar a las crías, y un control adecuado de la caza. Más recientemente también se ha considerado importante proveer alimento para las aves adultas y y hábitats para protegerlas de las aves rapaces. El abastecimiento del hábitat puede ser caro, pero en el Reino Unido los costos pueden recuperarse parcialmente mediante proyectos agro-medioambientales. El propietario de la tierra aún tiene que pagar por los servicios de los guardabosques. Desde 2003/2004, una parte del área de estudio de Sussex de la GWCT ha puesto en práctica estos principios de gestión ambiental, con la intención de restaurar la caza de la perdiz pardilla en esta zona del sur de Inglaterra. Los resultados han sido impresionantes, con un aumento de la densidad de parejas en primavera de 0,3/100 ha en 2003 hasta casi 20 parejas/100 ha en el 2010, en un área total de más de 10 km2. Durante los últimos dos años se ha practicado la caza de la perdiz pardilla y los propietarios de las tierras y sus equipos se han ganado el reconocimiento nacional por su labor conservacionista. Palabras clave: Perdiz pardilla, Perdix perdix, Control de predadores, Medidas agro–medioambientales, Recuperación de la población. Received: 9 III 12; Conditional acceptance: 24 IV 12; Final acceptance: 23 V 12 J. A. Ewald, G. R. Potts & N. J. Aebischer, Game & Wildlife Conservation Trust, Fordingbridge, Hampshire, SP6 1EF, UK. Corresponding author: J. A. Ewald. E–mail: jewald@gwct.org.uk ISSN: 1578–665X

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Introduction Across Europe, grey partridges have shown a long– term decline, amounting to 82% since 1980 (PECBMS, 2011). As a widespread farmland bird it is included in the European Farmland Bird Index (EFBI), a biodiversity indicator in the suite of EU Structural and Sustainable Development Indicators. In the UK the number of grey partridges has declined by over 90% since the mid–1960s (Risely et al., 2011), resulting in the species appearing on the Red List of Birds of Conservation Concern (Eaton et al., 2009). The grey partridge is also a priority species identified by the UK Biodiversity Action Plan (Anon., 1995) and is one of 19 species included in the Farmland Bird Indicator, one of the UK government biodiversity indicators for the natural environment (DEFRA, 2011). The partridge declines have continued despite the fact that the contributing factors in the UK and elsewhere are well understood (Potts, 1986; Aebischer, 1997; Aebischer et al., 2000). The management needed to reverse these declines includes various habitat improvements and the control of the predators of clutches and incubating hens. There is experimental verification of the effectiveness of providing nesting and brood–rearing habitats. For example, grey partridge brood sizes nearly doubled where 'conservation headlands’ (the outer 6 m of cereal fields selectively sprayed with pesticides) were used (Sotherton, 1991). Grassy mid–field strips ('beetle banks') are a means of providing appropriate mid–field nesting cover (Thomas et al., 1991; Sotherton, 1995). Many of these habitat improvements have been included in English agri–environment schemes. The success of these schemes in helping the grey partridge depends on implementing all the options that restore the multiple habitat requirements of the partridge (Ewald et al., 2010), including the legal predator control directed at reducing nest predation (Tapper et al., 1996). In this paper we describe how targeted management on one part of the long–running Sussex study in Southern England has resulted in the re–establishment of a wild grey partridge shoot, after a period of 48 years. Material and methods The Sussex study area is located on the Sussex Downs between the rivers Arun and Adur in West Sussex. The soil type is chiefly chalk rendzina, with clay caps on the higher ground. The land is mainly managed through arable farming, with cereals, oilseed rape/peas/linseed interspersed with grass fields and a few scattered patches of traditional chalk grassland. Since 1968, the Game & Wildlife Conservation Trust (GWCT) has monitored grey partridges using post– harvest stubble counts (Potts, 1986; Potts & Aebischer, 1991, 1995). We record the number of young in each covey, as well as the number of adult males and females. This allows us to calculate the autumn density of grey partridges on the area. Spring breeding density is calculated from these autumn counts, with each male and any single female representing a spring pair. Additional parameters used in the analysis of grey

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partridge demography are also calculated from these autumn counts. We concentrate here on the two that were targeted by the management taking place on the study area, the percentage of chicks that survive up to six weeks of age (chick survival rate) and the percentage of spring pairs that successfully produced a brood (brood production rate). Following Potts (1986): if the geometric mean brood size is less than 10,

(

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otherwise: Geometric mean brood size Chick survival rate = 13.84 From the chick survival rate the number of chicks that would have hatched can be calculated as: Number of young counted Chicks hatching = Chick survival rate Brood size at hatching is remarkably constant, at 13.84 (Potts, 1986) so: Chicks hatching Broods hatched = 13.84 Brood production rate =

Broods hatched Spring pairs

The calculation of brood production rate depends on how the calculated number of broods hatched compares to the actual number of spring pairs counted. In the occasional case where more broods are calculated to have hatched than there are spring pairs accounted for in the autumn then the number of broods is made equal to the number of spring pairs. Young counted Chick survival rate = 13.84 x Number of broods calculated to have hatched Throughout the course of the Sussex study the GWCT has also monitored farm practices and game management undertaken by the farmers. Changes in land use, including the cropping pattern, location of beetle banks, conservation headlands and other changes in boundaries are recorded on an annual basis. Management In 2003, one of the landowners within the study area set out to restore grey partridge numbers on an area of 220 ha, extending this to 1,052 ha in 2007, utilising all possible habitat and legal predation control measures. The remaining 22 km2 of the study area forms what is termed here the 'conventional area'. Habitat management included increases in nesting, brood–rearing and over–winter covers. Nesting habitat has been improved through the addition of 25 km of beetle banks and hedgerows. The beetle banks have been planted mainly with cock’s–foot (Dactylis glomerata) with the addition of hawthorn (Crataegus monogyna) and blackthorn (Prunus spinosa) bushes and some other species such as holly (Ilex aquifolium)

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Fig. 1. Spring pair density (A) and autumn density (B) of grey partridges on the Sussex study area. The dashed line indicates the situation on the 10 km2 area managed since 2003/2004; the black line is the remaining 22 km2 area. The shaded area indicates the period when management was taking place on the managed area. Fig. 1. Densidad primaveral (A) y otoñal (B) de parejas de perdiz pardilla en el área de estudio de Sussex. La línea punteada indica la situación en el área de 10 km2 gestionada desde 2003/2004; la línea continua negra es el área de 22 km2 restante. La zona sombreada indica el periodo durante el cual se estaba gestionando el área intervenida.

at intervals for cover. This additional nesting cover means that the managed area now has 8.2 km of nesting cover for each square kilometre. Brood–rearing cover has been expanded to nearly 9% of the area, with a total of 97 ha of conservation headlands put in place with no application of herbicides or insecticides and no fertiliser. In addition, there is no summer use of insecticides on cereal crops and one third of the conservation headlands are not harvested, providing some cover and seeds into the autumn where spring crops are to follow. As well as the winter cover provided through the unharvested conservation headlands, 2.1 km/km2 of wild bird cover strips were sown, incorporating kale

(Brassica oleracea), chicory (Cichorium intybus), millet (Panicum miliaceum) and canary grasses (Phalaris canariensis and P. arundinacea). These provide cover through the winter months as well as some food resources throughout the year. Cropping patterns have been adjusted to ensure that there is a different crop on either side of a field boundary or beetle bank. Where possible, fields are sown in autumn on one side of a field boundary with spring sowing on the other side, ensuring that on at least one side of a boundary there is vegetative cover at all times of the year. Feeders containing wheat are placed at intervals along the beetle banks and field boundaries, at a rate of two feeders for every pair of

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Chick survival rate (%)

100 80

Managed area Conventional area

60 40 20 0 1970 1975 1980 1985 1990 1995 2000 2005 2010 Year

Fig. 2. Brood production rate (A) and chick survival rate (B) of grey partridges on the Sussex study area. The dashed line indicates the situation on the 10–km2 area managed since 2003/2004; the black line is the remaining 22 km2 area. The shaded area indicates the period when management was taking place on the managed area. Fig. 2. Tasa de producción de crías (A) y tasa de supervivencia de pollos (B) de perdiz pardilla en el área de estudio de Sussex. La línea punteada indica la situación en el área de 10 km2 gestionada desde 2003/2004; la línea continua negra es el área de 22 km2 restante. La zona sombreada indica el periodo durante el cual se estaba gestionando el área intervenida.

grey partridges or 40/100 ha. Feeders are filled from October into June. Grain is provided through spring in the belief that it allows female grey partridges to leave the nest and quickly eat their fill, reducing the chance of nest and hen predation. Grit (1.5 mm) is also provided at each feeder. Some of the new habitat has been financed through the use of agri–environment options, particularly the provision of conservation headlands and beetle banks through the Higher Level Scheme (HLS, Natural England, 2010). Further to these habitat improvements, three gamekeepers are employed. From February to July, most

of their time is devoted to legal predation control targeted at reducing predation on grey partridge adults, eggs and chicks. This consists of controlling numbers of foxes (Vulpes vulpes), stoats (Mustela erminea), weasels (Mustela nivalis), rats (Rattus norvegicus), carrion crows (Corvus corone) and magpies (Pica pica). Methods include shooting with a rifle at night, stopped snares (with a breakaway for non–targets) and Larsen traps (specifically for corvid control). In autumn and winter, the gamekeepers are in charge of establishing the specially created habitats and provide the supplementary food.

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Number of broods hatched

250

200

2010

150 2009

100 2008

Managed area Conventional area

2007 2006 2005

100 150 200 Spring pairs

250

Fig. 3. The effect of breaking the density dependence between number of broods hatched and spring pairs through predation control is illustrated by the dashed line, with the number of broods hatched increasing almost in line with increases in the number of spring pairs. The continued density dependence in brood production on the conventional area is illustrated with the black lines, where the low brood production rate with higher spring pair density does not allow for a year–on–year increase in numbers. Fig. 3. Se ilustra el efecto de romper la dependencia de la densidad entre el número de nidadas empolladas y las parejas en primavera, debido al control de la depredación. Mediante la línea de puntos, con el número de nidadas empolladas aumenta con el incremento del número de parejas en primavera. La dependencia continuada de la densidad, de la producción de nidadas en las áreas convencionales, viene ilustrada por las líneas continuas. En ellas la baja tasa de producción de nidadas con una mayor densidad de parejas en primavera no tiene en cuenta el aumento de año en año de las cifras.

Results The breeding pair density on the conventional area increased marginally, from 0.9 pairs per km2 in 2004 to 2.4 pairs/km2 in 2010 (fig. 1A). Breeding density on the managed area increased from 5.2 pairs/km2 to 20.1 pairs/km2 in the same time frame, a 3.8 times increase and a density higher than that seen in the early 1970s. The autumn densities on the managed area now exceed those from the early days of the Sussex study, with densities nearly four times the highest seen in the 1970s (fig. 1B). On the managed area the increase in nesting habitat, combined with the control of predation at the nest, has resulted in brood production rates (77%) that are double what they were before the management began (38%; fig. 2A). On the conventional area, brood production rates remained unchanged, averaging 49% from 2004 to 2010 compared to 48% before 2004. Comparing chick survival over the managed area to the rest of the Sussex study area from 2004 to 2010 showed that the rate on the managed area was 58% on average, while on the conventional area it was 35% (fig. 2B). An increase in brood production rate is a feature of the predator control instigated on the managed area. Earlier work on the Sussex study and on other UK areas has

shown that k–factor nest loss increases with nest density, i.e. that it is density–dependent, and that this relationship with density is removed by predator control (fig. 4.2 in Potts, 1986). Therefore, where no gamekeeper is present, nest losses increase steeply as nesting density increases. When a gamekeeper is employed the relationship between nest losses and nest density disappears, allowing the density of spring pairs to build up, as has been the case on the managed area (fig. 3). Since 2009, the landowner undertaking the management has reinstated sustainable grey partridge shooting on the area. The bag amounted to 12% and 25% of the autumn stock in 2009 and 2010, respectively. The shooting revenue, combined with the income generated from the agri–environment payments, helps offset the cost of management. The landowner has indicated that these two income streams, combined with the enjoyment he gets from his own shooting, balance out his investment in grey partridge conservation. Discussion The turnaround in grey partridge numbers on this part of the Sussex study area is a testament to the hard work of the landowner and his team. In the space of

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just seven years, the grey partridge has gone from nearly extinct on this area to densities that support sustainable driven shooting. The hard work of the landowner and his team has not gone unnoticed. In 2010 they received the prestigious Purdey’s Gold Medal for Game and Conservation. When conferring the award the judges commented that the project was a shining example of how shooting and conservation could work together for the good of biodiversity. The landowner himself sees this project as a means of encouraging others, indicating that the funds currently available for agri–environmental work would allow others to put in place the habitat management required to turn around the fortunes of the grey partridge. How do the results reported here compare to other areas where specific partridge habitat management has been put in place? The 2010 spring pair density on the managed area matches the densities (18 pairs/km2) found on the GWCT’s demonstration project in Hertfordshire, where both habitat management and predator control were used (Aebischer & Ewald, 2010). It is double the density on the Salisbury plain experimental area, where only predator control was used to boost grey partridge numbers (Tapper et al., 1996). It does, however, fall far short of the densities (80 pairs/km2) locally reached in some hunting estates in northern France, where similar habitat management and predator control appropriate for the region is practised (Bourdouxhe, 2002; Bro et al., 2005). The English grey partridge model (Potts, 1986) shows that predation control approximately triples the equilibrium population level expected from increases in insect–rich habitat and nesting cover alone (Aebischer, 1991). At 20 pairs/km2, the results on the managed area are in line with what the model predicts for a fully managed shot population. Without the predation control, the expectation would be around 7 pairs/km2. The chick survival rates seen on the conventional portion of the Sussex study area are slightly higher than the long–term average reported for the post–decline era (32.3%; Potts & Aebischer, 1995), indicating that the situation on conventional sites has, at least, not deteriorated. Chick survival rates on the managed area are back to the level recorded before grey partridges began declining in number with the onset of herbicide use in cereals (48.6%; Potts & Aebischer, 1995). They surpass those of the original experiments that verified the ability of conservation headlands to provide chick–rearing habitat, resulting in increases in grey partridge chick survival (average brood size on the managed area is 8.5 chicks, compared to 6.4 on the original conservation headland areas: Rands, 1985). The cereal area covered by conservation headlands in the managed area exceeds the recommendations first made when conservation headlands were developed (6%; Boatman & Sotherton, 1988) and this surely explains the higher chick survival rates. How likely is it that the management described here could become widespread throughout England? Given the agri–environment options available in

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England, the habitats established in Sussex could all be incorporated into farming regimes through the use of the Higher Level Stewardship scheme. In the case of the GWCT’s Partridge Count Scheme (PCS), increases in chick survival rates on the area managed by PCS members are associated with their use of in–field agri–environment options that are designed to provide chick–rearing and nesting habitats, namely conservation headlands and beetle banks (Ewald et al., 2010). The area managed does not have to be of the order of 10 km2 and could be made up of ground owned by one landowner or several working together. However, our experience is that at least one dedicated gamekeeper is needed for every 400 ha. The Salisbury plain experiment has shown that, in the absence of habitat management, predation control alone on an area of between 4 to 5 km2 can increase densities to allow for a shootable surplus of grey partridges (an average of 23%; Tapper et al., 1996) but with numbers too low to provide revenue. Having a shootable surplus of partridges allows some return on a landowner’s dedication towards grey partridge conservation and ensures the long–term viability of the management. It is crucial that the landowner is motivated to carry on this work, whether through conservation interest or though shooting, and this is the driving force behind the success of the Sussex project. The message that the GWCT’s PCS is trying to convey to its members across England is that it is possible to restore a wild grey partridge shoot on a modern, productive arable farm, provided about 9% of the total area is given over to partridge management. The agri–environment options available to help with this have never been better and all that remains is for more farmers and landowners to take up the challenge and restore grey partridges on their ground. The English agri–environment programme is considered to be one of the most complex in Europe, with its system of options allowing for greater flexibility than might be the case in other European countries. This complexity can be used to good effect for wildlife generally. A review of EU–wide agri–environment options found that just under half of all rural development programmes had options directed at management for wildlife including actions specifically aimed at providing food, nesting and breeding areas (Keenleyside et al., 2011). Only a quarter of the programmes in the newer EU countries had such options. The results here indicate how useful these options can be in restoring grey partridge numbers and the need to include them in deliberations on the post–2013 Common Agricultural Policy. Acknowledgements We would like to thank all the landowners, farmers and gamekeepers within the Sussex study area. They have allowed us access to count partridges, collect invertebrates and cheerfully responded to requests for information for over 40 years. In the

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long run, they and land managers like them are responsible for the future of grey partridge conservation across Europe. References Aebischer, N. J., 1991. Sustainable yields: gamebirds as a harvestable resource. Gibier Faune Sauvage, 8: 335–351. – 1997. Gamebirds: management of the grey partridge in Britain. In: Conservation and the Use of Wildlife Resources: 131–151 (M. Bolton, Ed.). Chapman & Hall, London. Aebischer, N. J. & Ewald, J. A., 2010. Grey Partridge Perdix perdix in the UK: recovery status, set–aside and shooting. Ibis, 152: 530–542. – 2012. The grey partridge in the UK: popiulation status, research, policy and prospects. Animal Biodiversity and Conservation, 35.2: 353–362. Aebischer, N. J., Green, R. E. & Evans, A. D., 2000. From science to recovery: four case studies of how research has been translated into conservation action in the UK. In: Ecology and Conservation of Lowland Farmland Birds: 43–54 (N. J. Aebischer, A. D. Evans, P. V. Grice & J. A. Vickery, Eds.). British Ornithologists’ Union, Tring. Anon., 1995. Biodiversity: The UK Steering Group Report. Volume 2: Action Plans. Her Majesty’s Stationery Office, London. Boatman, N. D. & Sotherton, N. W., 1988. The agronomic consequences and costs of managing field margins for game and wildlife conservation. Aspects of Applied Biology, 17: 47–56. Bourdouxhe, L., 2002. Cent quintaux, cent perdreaux. Chasse et Nature, 94: 21–24. Bro, E., Reitz, F. & Landry, P., 2005. Grey partridge Perdix perdix population status in central northern France: spatial variability in density and 1994–2004 trend. Wildlife Biology, 11: 287–298. DEFRA (Department for Environment, Food and Rural Affairs), 2011. UK Biodiversity Indicators in Your Pocket 2011. www.jncc.defra.gov.uk/biyp. Eaton, M. A., Brown, A. F., Noble, D. G., Musgrove, A. J., Hearn, R., Aebischer, N. J., Gibbons, D. W., Evans, A. & Gregory, R. D., 2009. Birds of Conservation Concern 3: the population status of birds in the United Kingdom, Channel Islands and the Isle of Man. British Birds, 102: 296–341. Ewald, J. A., Kingdon, N. G. & Santin–Janin, H., 2009. The GWCT Partridge Count Scheme: a volunteer–based monitoring and conservation promotion scheme. In: Gamebird, 2006: 27–37 (S. B. Cederbaum, B. C. Faircloth, T. M. Terhune, J. J. Thompson & J. P. Carroll, Eds.). Warnell School of Forestry, Athens, USA. Ewald, J. A., Aebischer, N. J., Richardson, S. M., Grice, P. V. & Cooke, A. I., 2010. The effect of

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agri–environment schemes on grey partridges at the farm level in England. Agriculture, Ecosystems and Environment, 138: 55–63. Keenleyside, C., Allen, B., Hart, K., Menadue, H., Stefanova, V., Prazan, J., Herzon. I., Clement, T., Povellato, A., Maciejczak, M. & Boatman, N., 2011. Delivering environmental benefits through entry level agri–environment schemes in the EU. Report Prepared for DG Environment, Project ENV.B.1/ ETU/2010/0035. Inst. for European Environmental Policy, London. Natural England, 2010. Higher level Stewardship, Environmental Stewardship Handbook, Third Edition. PECBMS (Pan–European Common Bird Monitoring Scheme), 2011. Population Trends of Common European Breeding Birds 2011. CSO, Prague. Breeding Birds 2011. CSO, Prague. Potts, G. R., 1986. The Partridge: Pesticides, Predation and Conservation. Collins, London. – 2007. We Need More Managers and Better Theorists. In: Wildlife Science: Linking Ecological Theory and Management Applications (D. G. Hewitt & T. E. Fulbright, Eds.). CRC Press, London, UK. Potts, G. R. & Aebischer, N. J., 991. Modelling the population dynamics of the Grey Partridge: conservation and management. In: Bird Population Studies: Relevance to Conservation and Management: 373–390 (C. M. Perrins, J.–D. Lebreton & G. J. M. Hirons, Eds.). Editorial, ciudad edición. – 1995. Population dynamics of the Grey Partridge Perdix perdix 1793–1993: monitoring, modelling and management. Ibis, 137 (Supplement 1): S29–S37. Rands, M. R. W., 1985. Pesticide use on cereals and the survival of grey partridge chicks: a field experiment. Journal of Applied Ecology, 22: 49–54. Risely, K., Renwick, A. R., Dadam, D., Eaton, M. A., Johnston, A., Baillie, S. R., Musgrove, A. J. & Noble, D. G., 2011. The Breeding Bird Survey 2010. BTO Research Report 597. British Trust for Ornithology, Thetford. Sotherton, N. W., 1991. Conservation Headlands: a practical combination of intensive cereal farming and conservation. In: The Ecology of Temperate Cereal Fields: 373–397 (L. G. Firbank, N. Carter, J. F. Derbyshire & G. R. Potts, Eds.). Blackwell Scientific Publications, Oxford. – 1995. Beetle Banks–helping nature to control pests. Pesticide Outlook, 6: 13–17. Tapper, S. C., Potts, G. R. & Brockless, M. H., 1996. The effect of an experimental reduction in predation pressure on the breeding success and population density of grey partridges Perdix perdix. Journal of Applied Ecology, 33: 965–978. Thomas, M. B., Wratten, S. D. & Sotherton, N. W., 1991. Creation of island habitats in farmland to manipulate populations of beneficial arthropods – predator densities and emigration. Journal of Applied Ecology, 28: 906–917.

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Rock partridge (Alectoris graeca graeca) population density and trends in central Greece V. A. Bontzorlos, C. G. Vlachos, D. E. Bakaloudis, E. N. Chatzinikos, E. A. Dedousopoulou, D. K. Kiousis & C. Thomaides

Bontzorlos, V. A., Vlachos, C. G., Bakaloudis, D. E., Chatzinikos, E. N., Dedousopoulou, E. A., Kiousis, D. K. & Thomaides, C., 2012. Rock partridge (Alectoris graeca graeca) population density and trends in central Greece. Animal Biodiversity and Conservation, 35.2: 371–380. Abstract Rock partridge (Alectoris graeca graeca) population density and trends in central Greece.— The rock partridge is an emblematic species of the Greek avifauna and one of the most important game species in the country. The present study, which combined long term in–situ counts with distance sampling methodology in central Greece, indicated that the species’ population in Greece is the highest within its European distribution, in contrast to all prior considerations. Inter–annual trends suggested a stable rock partridge population both within hunting areas and wildlife refuges, whereas during summer, the species presented significantly higher densities in altitudes of more than 1,000 m, most probably due to the effect of predation at lower zones. The similarity of population structure between wildlife refuges and hunting zones along with the stable population trends demonstrate that rock partridge harvest in the country is sustainable. Key words: Rock partridge, Alectoris graeca graeca, Greece, Population trends, ANOVA models, Constrained ordination, Sustainable harvest. Resumen Densidad de población y tendencias de la perdiz griega oriental (Alectoris graeca graeca) en Grecia central.— La perdiz griega es una especie emblemática de la avifauna griega y una de las especies cinegéticas más importantes del país. En este estudio, en el que combinamos recuentos in situ a largo plazo con la metodología de muestreo a distancia, en Grecia central, indicó que la población de dicha especie en Grecia es la mayor de toda la zona de distribución europea, contrastando con todos los estudios anteriores. Las tendencias interanuales sugirieron la existencia de una población de perdiz griega estable tanto en los cotos de caza como en los refugios de fauna, mientras que en verano, esta especie presentaba densidades significativamente más altas a altitudes de más de 1.000 m, probablemente debido a los efectos de la depredación en las zonas inferiores. La similitud de la estructura de la población entre los refugios de fauna y las zonas de caza, junto a las tendencias poblacionales estables, demostraron que la caza de la perdiz griega en el país es sostenible. Palabras clave: Perdiz griega, Alectoris graeca graeca, Grecia, Tendencias poblacionales, Modelos ANOVA, Ordenación constreñida, Caza sostenible. Received: 23 XII 11; Conditional acceptance: 16 IV 12; Final acceptance: 24 V 12 Vasileios A. Bontzorlos, Hellenic Hunters’ Confederation, 8 Fokionos & Ermou Str., 10563, Athens, Greece.– Christos G. Vlachos, Lab. of Wildlife Management, Dept. of Forestry and Natural Environment, Aristotle Univ. of Thessaloniki, P. O. Box 262, GR–5400 Thessaloniki, Greece.– Dimitris E. Bakaloudis, Dept. of Forestry and Management of Natural Environment (Annex of Drama), Technological Education Inst. of Kavala, GR–66100 Drama, Greece.– Evangelos N. Chatzinikos, Eleni A. Dedousopoulou & Dimitrios K. Kiousis: Hunting Federation of Sterea Hellas, 8 Fokionos & Ermou Str., GR–10563 Athens, Greece.– Christos Thomaides, Dept of Forestry and Management of Natural Environment (Annex of Karpenisi), Technological Education Inst. of Lamia, GR–36100 Karpenisi, Greece. Corresponding author: Vasileios Bontzorlos. E–mail: vasilibon@gmail.com ISSN: 1578–665X

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Introduction The rock partridge is a Palearctic species with a limited geographical distribution in central and south Europe (Cattadori et al., 1999; Randi, 2006). Within this distribution range, the rock partridge occurs with four different subspecies: Alectoris graeca saxatilis can be found in discrete and often isolated populations in mountainous regions of various countries which share the Alps, such as the Italian Dolomitic Alps (Cattadori et al., 1999, 2003), the French Alps (Bernard–Laurent, 1991, 2000), the Austrian Alps (Bednar–Field et al., 2011), and the Dinaric Alps of the southwestern Balkans (Vogrin, 2001); Alectoris graeca whitakeri is strictly endemic and present only in the island of Sicily in Italy (Corso, 2010); Alectoris graeca orlandei has a distribution in the central and southern Appenines in Italy (Amici et al., 2009; Rippa et al., 2011); and finally, Alectoris graeca graeca which is distributed in the Balkan peninsula and mainly in southern Bulgaria (Dragoev, 1974; Nikolov & Spasov, 2005; Boev et al., 2007), Albania (Lucchini & Randi, 1998; Randi, 2006), some states of the Former Republic of Yugoslavia (Muzinic, 1995; Stevanovic et al., 2005), and large part of Greece. In Greece specifically, the species is encountered in all mountainous continental regions from eastern Macedonia to Peloponnesus including the Ionian islands of Lefkada and Kefalonia (Papaevangelou et al., 2001; Manios, 2002; Manios et al., 2002a, 2002b; Triantafyllidis et al., 2005, 2007) and the Aegean island of Euboea (Manios, 2002). It is present at altitudes higher than 400m with a few exceptions below that, whereas the lowest breeding altitude is recorded at 120m (Vavalekas et al., 1993). The rock partridge is one of the most important game species in the country. The hunting season runs from 1 X to 15 XII. Hunting is allowed only three days a week (Wednesday, Saturday and Sunday), with a daily bag limit of two birds per hunter. Despite its popularity as a game species, the only previous information on its population status in Greece is available in Papaevangelou et al. (2001). The main goals of this study, the first of its kind carried out in the country, were to: (i) estimate rock partridge population density and trends from a broad geographical area in central Greece, (ii) study population fluctuations, both inter–annually and intra–annually, (iii) study population variability between hunting and no hunting areas, (iv) define the effect of various environmental gradients on the variability of rock partridge density during the study years, and (v) determine population stability and sustainability of harvest. Material and methods Study area Sterea Hellas extends from the Ionian Sea in the west, to the Aegean Sea in the east, covering approximately 2,400,000 ha (fig. 1). Within this geographical region, the typical habitat of rock partridge, comprises a total of 402,996 ha which is our study area, according to

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GIS analysis and the CORINE Land Cover 2000 (EEA) Programme, as defined above the 400 m mark. The study area is approximately 17% of Sterea Hellas (fig. 2). Within these 402,996 ha where the species is currently found, 77,000 ha are wildlife refuges where hunting is not allowed (fig. 2), covering approximately 19% of the study area. The typical habitat of rock partridge in the Sterea Hellas region comprises 25% of the species’ total habitat at a national level (1,571,450 ha). Since typical rock partridge habitat extends throughout the region of Sterea Hellas and part of Euboea Island, it includes various habitat types, such as transitional zones between shrubs and woodland (68%), moors, heathland and sclerophylous vegetation (5.6%), natural grasslands (17.5%), sparsely vegetated areas (6.4%), pastures (1%) and bare rocky areas (0.8%). Field methods Monitoring the rock partridge population in central Greece began in 2005 as a continuous research programme financed by the Hunting Federation of Sterea Hellas. The data presented here are for the first seven years of the programme (2005–2011). A total of 65 line transects were located within wildlife refuges and 80 line transects within hunting areas, spread throughout the study area, both in latitude and altitude. The mean length of each transect was 2.5 km, ranging from 1.9 to 3.3 km (95% CI: 2,405–2,632 m). Line transects did not vary during the seven years of study. Monitoring was carried out according to the line transect method (Buckland, 2001) by gamekeepers using pointing dogs (Sara, 1989; Cattadori et al., 2003; Besnard et al., 2010). Each year two counts were conducted on each line transect. One count was conducted during the last two weeks of March because in that period the formation of rock partridge pairs is complete and the breeding season is about to begin. The second count was conducted during the last two weeks of August because in that period the breeding season is over and it is the best time to distinguish and record the number of younglings (Manios, 2002). During the counts, one gamekeeper walked the line transect and two others covered the area to the left and right of the line with pointing dogs. When partridges were flushed, the perpendicular distance of individuals from the line transect was recorded. Along with the number of individuals flushed each time, we also recorded number of pairs during March and the number of individuals per flock during August (Buckland, 2001). In addition, number of adults and juveniles was also recorded during counts in late summer. During the seven years of the study, a total census of 1,275 km was done within hunting areas and 1,050 km within wildlife refuges. Vegetation type, vegetation cover and grazing intensity were also recorded as environmental variable categories. Statistical analyses The species’ population density was calculated using Distance 6.0 Release 2 (Buckland et al., 2008; Fewster et al., 2009; Thomas et al., 2010). Rock partridge

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density was calculated for each year, season, area and altitudinal zone. In order to test for variability in rock partridge population trends between years and between seasons, and also between wildlife refuges and hunting areas, a three–way full factorial ANOVA was constructed. The first factor of predictor variables was the year (2005–2011), the second factor was the season (early spring vs. late summer) and the third factor was the area (wildlife refuges vs. hunting areas). Moreover, in order to explore any possible altitudinal effect on rock partridge density, another two–way full factorial ANOVA model was constructed with the first factor of predictor variables being the season and the second factor the altitude. Altitude was divided into two zones, a low zone from 400 m to 1,000 m and a high zone from 1,001 m to 2,000 m. Altitude was not included in the first ANOVA model as the fourth factor of predictor variables because in order to run such a model at least two values of rock partridge density (ind/ha) should be produced for each combination of the four factors’ levels: year (2005–2011), season (spring–summer), area (wildlife refuges–hunting areas) and altitude (low zone–high zone). Density values, however, were calculated using Distance software which has a minimum threshold of 40 species encounters during field counts, for each level combination, in order to produce results (Buckland et al., 2001, 2008). This limitation occurred on certain occasions when building a four–way ANOVA, and thus the effect of altitude and its interaction with season was explored in the present study using a second ANOVA model.

N W

E S

373

1:4,000,000

N W

E S Greece Region of Sterea Hellas

Fig. 1. Map of Greece showing the Sterea Hellas region. Fig. 1. Mapa de Grecia mostrando la región de Sterea Hellas.

Region of Stenea Helllas Hunting zones Wildlife refuges

1:1,300,000

Fig. 2. Map of the Sterea Hellas region (central Greece) indicating the total typical habitat of rock partridge. Hunting zones in the total study area are indicated in black and wildlife refuges in grey. Fig. 2. Mapa de la región de Sterea Hellas (Grecia central) indicando el total del hábitat típico de perdiz griega. Las zonas de caza en el área total de estudio se indican en negro, y los refugios de fauna en gris.

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Table 1. Mean rock partridge density (ind/ha) in the Sterea Hellas region (central Greece) from 2005 to 2011, according to season, area and altitudinal zone. Tabla 1. Densidad media de perdiz griega (ind/ha) en la región de Sterea Hellas (Grecia central) del 2005 al 2011, según la estación, el área y la franja altitudinal.

Wildlife refuge

Hunting zones

Low altitude 400–1,000 m

Spring (SD)

0.185 (0,048)

0.123 (0,026)

0.145 (0,038)

0.164 (0,058)

Summer (SD)

0.468 (0,094)

0.367 (0,116)

0.348 (0,106)

0.486 (0,079)

The effect of habitat type, vegetation cover, grazing density and altitude upon the rock partridge density during spring and summer was also analyzed using a multivariate approach through constrained ordination and Generalized Linear Models (Leps & Smilauer, 2003). A 'response' variables matrix was constructed at first, including rock partridge absolute count numbers in spring and summer as recorded in each line transect and their repeated results each year of the study. Then, a similar 'predictor' variables matrix was constructed including habitat type, vegetation cover, grazing density and altitude as recorded in each line transect, season and year. Except altitude, which is a continuous variable, the remaining environmental variables were recorded in a categorical form. Thus they were obligatory transformed through fuzzy coding in order to be expressed with binomial values (absence: 0, presence: 1), so that they could be included in the multivariate analysis (Leps & Smilauer, 2003). Once both matrices were introduced in the software, a Detrended Correspondence Analysis was firstly applied only on the 'response' matrix. This type of indirect analysis considers only the variability of response variables and calculates length gradient values which are actually measurements of beta diversity in community composition. According to their value (less than 3 or more than 4), these results indicate the kind of multivariate approach —linear or unimodal analyses— to be followed, respectively. The indicated analysis was then applied to both matrices, in order to produce a constrained ordination which represents the variability in rock partridge population composition that can be explained by the measured predictor variables. In constrained ordination specifically, the produced axes are weighed sums of the predictor variables, and thus these methods of direct gradient analysis resemble a model of multivariate multiple regression. The significance of the model is then tested with Monte Carlo simulations, and if significant, various hypotheses can be explored with the use of GLMs and the criterion of Akaike (AIC). The percentage of juveniles with respect to the total number of individuals during late summer in each monitoring effort was also calculated. Differences between wildlife refuges and hunting areas were explored with a one–way ANOVA model in order to test whether the reproductive outcome varies among them. In order to apply all the ANOVA models, absolute

High altitude 1,000–2,000 m

counts and density values were log–transformed to meet the assumptions of the analysis. Results Based on the species’ mean density (ind/ha) from all the years of the study (table 1) the rock partridge population in Sterea Hellas, was estimated at a mean of 31,000 breeding pairs in early spring (95% CI: 28,052 to 34,358).

Table 2. Results of full factorial model for three– way ANOVA on rock partridge population density in the Sterea Hellas region (central Greece) from 2005 to 2011. The first factor in the model is the year of the study, the second factor is the season (early spring–late summer) and the third factor is the area (wildlife refuges–hunting areas): Df. Degrees of freedom. Tabla 2. Resultados del modelo factorial completo para una ANOVA de tres factores sobre la densidad de población de perdiz griega en la región de Sterea Hellas (Grecia central) del 2005 al 2011. El primer factor del modelo es el año de estudio, el segundo es la estación del año (principios primavera–finales del verano), y el tercero es la zona (refugios de fauna–áreas de caza): Df. Grados de libertad. Df

Year

6,28

0.336

0.911

Season

1,28 116.635 < 10–6

Area

1,28

11.280

0.002

Year*Season

6,28

0.146

0.988

Year*Area

6,28

0.508

0.796

Season*Area

1,28

0.273

0.604

Year*Season*Area

6,28

0.288

0.937

Animal Biodiversity and Conservation 35.2 (2012)

Wildlife refuge Hunting area

0.25

Wildlife refuge Hunting area

0.20 0.15 0.10

Spring

2011

2010

2009

2008

2005

2011

2010

2009

2008

2007

2006

2005

0.00

2007

0.05

2006

Rock partridge density log 10(x+1)

0.30

375

Summer

Fig. 3. Variability of rock partridge density from 2005 to 2011, in spring and summer, within wildlife refuges and hunting areas in the Sterea Hellas region (central Greece). Fig. 3. Variabilidad de la densidad de perdiz griega desde el año 2005 al 2011, en primavera y en verano, dentro de los refugios de fauna y en las zonas de caza de la región de Sterea Hellas (Grecia central).

The year factor played no significant role in the inter–annual variability of the rock partridge density in the study area (table 2). Similarly, no significant variance occurred in rock partridge inter–annual population trends in different seasons, or within hunting areas and wildlife refuges (table 2). On the other hand, we found a strong intra–annual seasonal effect due to higher densities during summer both within hunting areas and wildlife refuges (table 2, fig. 3). Finally, there was also a significant difference in rock partridge density; it derived from higher densities within wildlife refuges than in hunting areas, which was not as strong as the seasonal effect but it was constant during all the years of the study, both during spring and summer (table 2, fig. 3). Rock partridge density also increased significantly along the altitudinal gradient (table 3). Nonetheless, the significant interaction between seasonal and altitudinal effects showed that altitude had a significant effect on the increase in the species’ density only during summer, because during spring, rock partridge density presented no difference between low and high altitudinal zones (fig. 4). The positive effect of altitudinal gradient upon the rock partridge density was also verified by the constrained ordination. Detrendend Correspondence Analysis (DCA) on the “response” variables dataset produced a length gradient of less than 3 for the first axis, indicating that linear methods should be used in continuation and specifically an RDA (Redundancy Analysis). Direct gradient analysis through RDA upon both predictor and response variables’ datasets produced a significant model (F–ratio = 5,953, p = 0.002), from which the first

produced constrained canonical axis explained 99% of the variability in rock partridge density (table 4). Forward selection results of the produced multivariate model, according to both marginal and conditional effects, indicated 'Altitude' as the most important variable. In the two–dimensional representation of rock partridge variability during spring and summer and its position in ordinational space, altitude defined the horizontal axis upon which the vectors of the rock partridge density increased, and specifically that of summer. Environmental variables 'Low grazing', 'Shrubs' and 'Phrygana' occupied the 4th quadrant of the graph, 'High grazing' and 'Medium grazing' occupied the upper and lower parts of the bi–plot and 'Subalpine' occurred alone in the second quadrant (fig. 5). Rock partridge population structure in central Greece was similar both within hunting zones and wildlife refuges (one–way ANOVA: F1,12 = 0.020; p = 0.890). The average ratio of juveniles to adults within hunting areas was 1.6 (range 1.2–1.9; SD 0.29) and within wildlife refuges it was 1.5 (range 1.2–1.8; SD 0.22). Discussion The rock partridge population in Greece has been estimated at between 7,000 and 13,000 breeding pairs by Handrinos & Akriotis (1997) and Handrinos & Papoulia (2004), with a claim of being even lower (Handrinos & Katsadorakis, 2009). These same authors also consider the species to be extremely rare and definitely declining and disappearing from most parts of Greece. The authors state, however, that

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Table 3. Results of the full factorial model for two–way ANOVA on rock partridge population density in the Sterea Hellas region (central Greece) from 2005 to 2011. The first factor is the season of the counts (early spring–late summer) and the second factor is the altitude (low altitude: 400–1000 m; high altitude: 1,001–2,000 m): Df. Degrees of freedom. Tabla 3. Resultados del modelo factorial completo para una ANOVA de dos factores sobre la densidad de población de perdiz griega en la región de Sterea Hellas (Grecia central) del 2005 al 2011. El primer factor es la estación de los recuentos (principios primavera–finales del verano) y el segundo es la altitud (baja altitud: 400–1.000 m; gran altitud: 1.001–2.000 m): Df. Grados de libertad.

Season

1,52

183.817

< 10–6

Altitude

1,52

14.579

< 10–3

Season*Altitude

1,52

7.857

0.007

these estimations are not based on specific scientific data on rock partridge populations (Handrinos & Katsadorakis, 2009). These estimations were based on an older publication of Papaevangelou et al. (2001) in which rock partridge population in Greece was estimated at between 7,000 and 13,000 pairs, without any clear indication of the methodological approach used to determine these numbers, or any reference concerning the mathematical procedure used to calculate the total number of breeding pairs at a national level.

In contrast, according to our study, the first to be carried out in such an extended study area and on such a long term basis, Greece holds the highest Alectoris graeca graeca population in the Balkans. Moreover, Greece also holds the highest population among all other countries within the species’ European distribution range. Sterea Hellas alone holds an estimated total of 31,000 breeding pairs. Furthermore, attempting a simple extrapolation of the species’ mean density on the total rock partridge typical habitat in the country (1,571,450 ha), the breeding population of the species in Greece is potentially estimated at approximately 121,000 pairs (95% CI: 109,338 to 133,979), which is much higher than any previously published estimations. Although in certain regions of the country the rock partridge density may vary from the one calculated in Sterea Hellas (table 1), the present findings provide a solid base for an estimation of national rock partridge population. According to our data, the minimum estimated number of rock partridge breeding pairs in Greece is 104,000, which is much higher from the species’ maximum European breeding population, as it was previously estimated at 78,000 individuals (Burfield & Bommel, 2004). According to the results of Project ARTEMIS, since 2007 when the bag limit was set at two birds per hunter and outing, harvest of rock partridge in the Sterea Hellas region is stable and calculated (Thomaides et al., 2011) at a mean of 26,000 individuals per year (95% CI: 22,954 to 29,423). This further attests to harvest sustainability and population stability of the species. Apart from the fact that our data give a completely different picture concerning the rock partridge population status in Greece, they also demonstrate that there has been no declining trend for rock partridge in the seven years of the study. The species’ populations are stable both within wildlife refuges and hunting zones, as well as during the monitoring seasons, early spring and late summer (table 2, fig. 3). The constant difference in rock partridge density between hunting zones and

Table 4. Results of constrained ordination analysis (RDA) for the rock partridge density dataset during spring and summer, and for the environmental variables dataset in the Sterea Hellas region (central Greece) from 2005 to 2011. Tabla 4. Resultados del análisis de ordenación constreñida (RDA) del conjunto de datos sobre la densidad de perdiz griega durante la primavera y el verano, y el conjunto de datos de las variables ambientales de la región de Sterea Hellas (Grecia central) del 2005 al 2011.

Axes

Eigenvalues

0.075

0.001

0.634

0.289

Species–environment correlations

0.330

0.056

0.0

7.6

71.1

100.0

0.0

Cumulative percentage variance of species data

7.5

Cumulative percentage variance of species–environment relation 98.7

Animal Biodiversity and Conservation 35.2 (2012)

0.26

Rock partridge density log 10 (x + 1)

0.24 0.22

377

High altitudinal zone (1,000–2,000 m) Low altitudinal zone (400–1,000 m)

0.20 0.18 0.16 0.14 0.12 0.10 0.08 0.06 0.04

0.02

Spring

Summer

Fig. 4. Variability of rock partridge density from 2005 to 2011, in low and high altitudinal zones, during spring and summer, in the Sterea Hellas region (central Greece). Fig. 4. Variabilidad de la densidad de perdiz griega de 2005 a 2011, en zonas altitudinales bajas y altas, durante la primavera y el verano, en la región de Sterea Hellas (Grecia central).

wildlife refuges is only contradicted during the summer of 2010, when the species’ density within wildlife refuges was at its lowest (fig. 3). Although the species’ interannual population trends are stable according to our results, this finding could be a first indication for cyclic fluctuations of rock partridge in Greece. Cattadori et al. (1999) first recorded rock partridge fluctuations in the Dolomitic Alps, after analysing harvest data over 40 years, defining a cyclic period of four to seven years. Of course, more than seven years of data need to be taken into consideration to reach a safe conclusion for Greece. However, the fact that rock partridge showed their lowest density value in our study after five years of summer counts, and that this was reversed the following year, probably points to a similar conclusion. It is also important that this phenomenon was observed only in wildlife refuges and not within hunting zones. Nevertheles, more in depth research concerning demographic parameters should be conducted to give specific answers to this issue. Although hunting zones and wildlife refuges are adjacent, as indicated in figure 2, the strong territorial nature of rock partridge and the species’ small range movements lend meaning to statistical comparisons between the two areas (Manios, 2002; Manios et al., 2003). In addition, line transects were not placed near the borders of hunting zones and wildlife refuges. Moreover, wildlife refuges in Greece include very large areas of various habitat types or biogegographical units such as whole mountains, and thus we can safely support that there are no rock partridge movements between refuges and hunting areas.

The significant effect of season upon rock partridge density (table 2) was expected, because reproductive output contributed towards a significant increase in rock partridge density, both within hunting zones and wildlife refuges (fig. 3). On the other hand, the significant effect of altitude on the increase of rock partridge density during summer and specifically on the higher altitudinal zones (table 3, fig. 4) is probably due to the effect of predation upon rock partridge nests and nestlings at lower altitudes. As recorded by Manios (2002) in Greece, 72% of the located rock partridge nests were destroyed, mainly by beech marten (Martes foina) and weasel (Mustela nivalis) in lower altitudes, probably accounting for the significantly higher density values in the 1,000m to 2,000 m zone in summer (fig. 4). Lower altitudes sustain more densely vegetated ecosystems with higher canopy cover, habitats that are ideal for predators such as marten and weasel (Spencer et al., 1983; Cavallini & Lovari, 1991; Clevenger, 1994; Sachhi & Meriggi, 1995; Lucherini et al., 1995), whereas at higher altitudes where vegetation is sparse, open areas sustain poorer assemblages in respect to these two species and thus predation upon rock partridge nests and nestlings decreases significantly. It is also possible that rock partridges move in accordance with climatic conditions during each season. In the French Alps, it was recorded that the species tried to reach its preferred habitats during summer, these being habitats located at higher altitudes. Specifically, rock partridges moved several kilometres away from their breeding sites during winter in order to avoid heavy snowfall,

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High grazing Low grazing

Shrubs

Phyrgana Rock partridge summer density Rock partridge spring density

Altitude

Subalpine Medium grazing

Fig. 5. Variability in rock partridge density from 2005 to 2011 in the Sterea Hellas region (central Greece) during spring and summer (response variables) explained by environmental gradient variability (predictor variables), presented in two–dimensional space produced by RDA. Triangles represent qualitative environmental variables, the dashed vector represents the increase of altitude which is the only continuous variable, and black vectors indicate increase density of rock partridge in spring and summer (ind/ha). The proximity of a black vector towards a triangle or the direction of the dashed vector indicates density increase upon this gradient, whereas distance or opposite direction indicate density decrease and negative gradient effect, respectively. Fig. 5. Variabilidad de la densidad de perdiz griega del 2005 al 2011 en la región de Sterea Hellas (Grecia central) durante la primavera y el verano (variables de respuesta) explicada por la variabilidad del gradiente ambiental (variables predictivas), presentada en un espacio bidimensional producido mediante RDA. Los triángulos representan las variables ambientales cualitativas, el vector punteado representa el aumento de altitud, la única variable continua, y los vectores negros indican el aumento de la densidad de perdiz griega durante la primavera y el verano (ind/ha). La proximidad de un vector negro a un triángulo, o la dirección del vector punteado indican un aumento de la densidad hacia el gradiente, mientras que la distancia o la dirección opuesta indican un descenso de la densidad y un efecto de gradiente negativo, respectivamente.

and returned to the same sites during summer when these became accessible (Bernard–Laurent, 1991). The altitudinal effect on rock partridge density in summer is also verified from the constrained ordination and the significance of the model that is shown in the ordination bi–plot (fig. 5). The large vector length of the environmental variable 'Altitude' indicates its importance in explaining rock partridge variability in the model. Moreover, since the vector is almost parallel to the horizontal axis, it is the environmental variable which actually defines it. The importance of altitude in the model is also confirmed from RDA results, which indicate that the first produced constrained axis, which is defined by 'Altitude', explains almost 99% of rock partridge variability in the model. In addition, from the response variables’ dataset only the rock partridge

density during summer has a large vector length and is parallel to the horizontal axis, indicating that altitude mainly affects the species’ density during summer, which increases along the gradient. The other environmental variables explain only the remaining 1% of the species’ variability. Specifically, 'Low grazing', 'Shrubs' and 'Phrygana' are clustered in the fourth quadrant of the graph, in the opposite direction of the increasing density of rock partridge, possibly demonstrating a negative effect. This could be an indication that the rock partridge in Greece prefers early succession ecosystems with open areas, but with just 1% of explained variance such an argument is not fortified. Nonetheless, such behavior has been recorded previously in Greece by Manios (2002) and in Italy by Rippa et al. (2011), where higher grazing intensity

Animal Biodiversity and Conservation 35.2 (2012)

created more open habitats which proved to be more suitable for the species. Finally, the fact that inter–annual rock partridge density trends demonstrate a stable population, both within hunting areas and wildlife refuges, as well as the fact that the species’ population structure is almost identical in both areas through the years, suggest that the species’ harvest is sustainable and the reproductive output compensates for the hunting take. The average productivity ratio (juveniles/adults) as calculated in central Greece both in hunting zones and wildlife refuges, is similar to that provided by Bernard–Laurent (1994) and Sara (1989). According to Sara (1989), it is characterized as a medium productivity, but in our study higher densities per land unit are demonstrated (table 1). In conclusion, since hunting appears to have no negative effect on the species’ population, the limiting factors for rock partridge population, as recorded by other authors, are: genetic hybridization (Triantafyllidis et al., 2005, 2007), predation (Manios, 2002; Vavalekas et al., 1993), the abandonment of traditional agriculture and livestock practices leading to increased canopy cover in mountainous areas (Papaevangelou et al., 2001; Manios, 2002; Rippa et al., 2011), and the effect of parasites (Manios et al., 2002b; Rosa et al., 2011). Acknowledgements We would like to thank the gamekeepers at the Hunting Federation of Sterea Hellas for conducting the monitoring process in the field. We would also like to thank the Hellenic Hunters Confederation for providing the harvest data. This study was financed by the Hunting Federation of Sterea Hellas. References Amici, A., Pelorosso, R., Serrani, F. & Boccia, L., 2009. A nesting site suitability model for rock partridge (Alectoris graeca) in the Apennine Mountains using logistic regression. Italian Journal of Animal Science, 8(Supplement 2): 751–753. Bednar–Friedl, B., Behrens, D. A. & Getzner, M., 2011. Optimal dynamic control of visitors and endangered species in a national park. Environmental and Resource Economics, 28 (September 2011): 1–22. Bernard–Laurent, A., 1991. Migrant rock partridges (Alectoris graeca saxatilis) in the southern French Alps. Journal of Ornithology, 132: 220–223. – 1994. Statut, evolution et facteurs limitant les populations de perdrix bartavelle (Alectoris graeca): Synthese bibliographique. Game and Wildlife, 11 (Hors serie Tome 1): 267–307. – 2000. Vulnerability of an alpine population of rock partridge (Alectoris graeca saxatilis) to climatic events: Evaluation with deterministic and stochastic models. Game and Wildlife Science, 17: 63–79. Besnard, A., Novoa, C. & Jimenez, O., 2010. Hunting impact on the population dynamics of Pyrenean grey partridge Perdix perdix hispaniensis. Wildlife Biology, 16: 135–143.

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Boev, Z., Milchev, B. & Popov, V., 2007. Fauna, zoogeography, and ecology of birds in Bulgaria. In: Biogeography and Ecology of Bulgaria, Monographiae, 82: 39–78 (V. Fet & A. Popov, Eds.). Springer–Verlag, New York Inc. Buckland, S. T., Anderson, D. R., Burnham, K. P., Laake, J. L., Borchers, D. L. & Thomas, L., 2001. Introduction to distance sampling–estimating abundance of biological populations. Oxford Univ. Press, Oxford. Buckland, S. T., Marsden, S. & Green, R. E., 2008. Estimating bird abundance: making methods work. Bird Conservation International, 18: 91–108. Burfield, I. & Bommel, F., 2004. Birds in Europe. Population estimates, trends and conservation status. Birdlife Conservation Series, No 12. Birdlife International, Cambridge. Cattadori, I. M., Hudson, P. J., Merler, S. & Rizzoli, A., 1999. Synchrony, scale and temporal dynamics of rock partridge (Alectoris graeca saxatilis) populations in the dolomites. Journal of Animal Ecology, 68: 540–549. Cattadori, I. M., Ranci–Ortigosa, G., Gatto, M. & Hudson, P. J., 2003. Is the rock partridge Alectoris graeca saxatilis threatened in the Dolomitic Alps? Animal Conservation, 6: 71–81. Cavallini, P. & Lovari, S., 1991. Environmental factors influencing the use of habitat in the red fox, Vulpes vulpes. Journal of Zoology, 223: 323–339. Clevenger, A. P., 1994. Habitat characteristics of Eurasian pine martens Martes martes in an insular Mediterranean environment. Ecography, 17: 257–263. Corso, A., 2010. Sicilian rock partridge: Identification and taxonomy. Dutch Birding, 32: 79–96. Dragoev, V., 1974. On the population of the rock partridge (Alectoris graeca Meisner) in Bulgaria and methods of census. Acta Ornithologica, 14: 251–255. Fewster, R. M., Buckland, S. T., Burnham, K. P., Borchers, D. L., Jupp, P. E., Laake J. L. & Thomas, L., 2009. Estimating the encounter rate variance in distance sampling. Biometrics, 65: 225–236. Handrinos, G. & Akriotis, T., 1997. The Birds of Greece. Christopher Helm, London. Handrinos, G. & Katsadorakis, G., 2009. Alectoris graeca Rock partridge. In: The Red Book of threatened animals of Greece: 290–291 (A. Legakis & P. Maragou, Eds.). Hellenic Zoological Society, Athens. Handrinos, G. & Papoulia, S., 2004. Alectoris graeca Rock partridge. In: Birds in Europe. Population estimates, trends and conservation status. Birdlife Conservation Series No 12: 95 (I. Burfield & F. van Bommel, Eds.). Information Press, Oxford, UK. Leps, J. & Smilauer, P., 2003. Multivariate analysis of ecological data using CANOCO. Cambridge Univ. Press. Lucherini, M., Lovari, S. & Crema, G., 1995. Habitat use and ranging behaviour of the red fox (Vulpes vulpes) in a Mediterranean rural area: is shelter availability a key factor? Journal of Zoology, 237: 577–591.

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Lucchini, V. & Randi E., 1998. Mitochondrial DNA sequence variation and phylogeographical structure of rock partridge (Alectoris graeca) populations. Heredity, 81: 528–536. Manios, N., 2002. The ecology of the rock partridge Alectoris graeca graeca in Epirus and Fokida (Greece). Ph. D. Thesis, Aristotle Univ. of Thessaloniki. Manios, N., Alexiou, B., Chatzinikos, E., Papageorgiou, N. & Tsachalidis, E., 2002a. Naturally marked individuals of rock partridge (Alectoris graeca graeca) in Greece. European Journal of Wildlife Research, 48 (Supplement 1): 373–377. Manios, N., Papazahariadou, M., Frydas, S., Papageorgiou, N., Tsachalidis, E. & Georgopoulou, J., 2002b. Tetrathyridium as a mortality factor of rock partridge (Alectoris graeca graeca) in central Greece. European Journal of Wildlife Research, 48 (Supplement 1): 378–382. Manios, N., Aleksiou, V., Chatzinikos, E. & Papageorgiou, N., 2003. Homerange and seasonal movements of rock partridge (Alectoris graeca graeca) in Greece. In: Integrating Wildlife with People. Abstracts and contributing authors of the XXVI Congress of the International Union of Game Biologists: 121 (J. V. Vingada, Ed.). Braga, Portugal. Muzinic, J., 1995. The state of bird and nature protection in Croatia. The Environmentalist, 15: 188–195. Nikolov, S. C. & Spasov, S. D., 2005. Frequency, density and numbers of some breeding birds in the south part of Kresna George (SW Bulgaria). Acrocephalus, 26: 273–282. Papaevangelou, E., Thomaides, C., Handrinos, G. & Haralambides, A., 2001. Status of partridge (Alectoris and Perdix) species in Greece. Game and Wildlife Science, 18: 253–260. Randi, E., 2006. Evolutionary and conservation genetics of the rock partridge Alectoris graeca. Acta Zoologica Sinica, 52 (Supplement): 370–374. Rippa, D., Maselli, V., Soppelsa, O. & Fulgione, D., 2011. The impact of agro–pastoral abandonment on the rock partridge Alectoris graeca in the Appenines. Ibis, 153: 721–734. Rosa, R., Bolzoni, L., Rosso, F., Pugliese, A., Hudson, P. J. & Rizzoli, A., 2011. Effect of Ascaridia compar infection on rock partridge population dynamics: empirical and theoretical investigations. Oikos, 120: 1557–1567. Sacchi, O. & Meriggi, A., 1995. Habitat requirements

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of the stone marten (Martes foina) on the Tyrrhenian slopes of the Northern Appenines. Hystrix, 7: 99–104. Sara, M., 1989. Density and biology of the rock partridge (Alectoris graeca whitakeri) in Sicily (Italy). Italian Journal of Zoology, 56: 151–157. Spencer, W. D., Barrett, R. H. & Zielinski, W. J., 1983. Marten habitat preferences in the northern Sierra Nevada. Journal of Wildlife Management, 47: 1181–1186. Stevanovic, S., Pavlovic, I. & Stevanovic, D., 2005. Rock partridge (Alectoris graeca L.). Zivinarstvo, 40: 32–33. Thomaides, C., Logothetis, G., Karabatzakis, T. & Christoforidou, G., 2011. Project ARTEMIS: The game statistics survey in Greece–Monitoring game populations and game harvest in Greece during the years 1995/1996 thru 2009–2010. Hellenic Hunters Confederation, Athens. Thomas, L., Buckland, S. T., Rexstad, E. A., Laake, J. L., Strindberg, S., Hedley, S. L., Bishop, J. R. B., Marques, T. A. & Burnham, K. P., 2010. Distance software: design and analysis of distance sampling surveys for estimating population size. Journal of Applied Ecology, 47: 5–14. Triantafyllidis, A., Alexandri, P., Ververis, A., Tilaveridou, K., Chatzinikos, E., Manios, N., Papageorgiou, N., Triantafyllidis, C., 2007. Genetic structure and hybridization of Greek partridges Alectoris graeca and Alectoris chukar based on microsatellite DNA analysis. In: Book of Abstracts of the XXVIII Congress of the International Union of Game Biologists: 132 (K. Sjoberg & T. Rooke, Eds.). Sodra Tornet Kommunikation, Uppsala. Triantafyllidis, A., Karatzas, D., Georgiadou, A., Drikos, I., Andreakou, E., Lappa, M., Chatzinikos, E., Manios, N., Papageorgiou, N., Triantafyllidis, C., 2005. Genetic identification of Greek partridges Alectoris graeca and Alectoris chukar. In: Extended Abstracts of the XXVII Congress of the International Union of Game Biologists: 188–190 (K. Pohlmeyer, Ed.). Edition Natur Life, Hamburg. Vavalekas, K., Thomaides, C., Papaevangelou E. & Papageorgiou, N., 1993. Nesting biology of the rock partridge Alectoris graeca graeca in northern Greece. Acta Ornithologica, 28: 97–101. Vogrin, M., 2001. Overview of Slovenian ornithofauna. Acta Zoologica Lituanica, 2: 20–24.

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Restoration of a sustainable wild grey partridge shoot in eastern England R. A. H. Draycott

Draycott, R. A.H., 2012. Restoration of a sustainable wild grey partridge shoot in eastern England. Animal Biodiversity and Conservation, 35.2: 381–386. Abstract Restoration of a sustainable wild grey partridge shoot in eastern England.— Eastern England has been a stronghold for grey partridges Perdix perdix, but in common with the rest of Britain, numbers declined from the 1950s onwards. Partridges within a 40 km2 study area in the county of Norfolk have been monitored in conjunction with the Game and Wildlife Conservation Trust (GWCT) since the 1950s. Since 2001 a programme of habitat creation, supplementary feeding and predation control was undertaken by the landowner, farmers and gamekeepers to restore partridges. Numbers increased from 4.7 pairs/km2 in March 2001 to 54 pairs/km2 in March 2011. These densities are comparable with those before the national decline in grey partridge stock. In the last three winters, between 13 and 74 birds/km2 were harvested and spring stocks continue to increase. Key words: Grey partridge, Habitat creation, Supplementary feeding, Predation control, Shoot. Resumen Recuperación de un coto de caza sostenible de perdiz pardilla en el este de Inglaterra.— El este de Inglaterra ha sido un baluarte de la perdiz pardilla, Perdix perdix, pero al igual que en el resto de Gran Bretaña, sus efectivos están disminuyendo desde los años cincuenta. Desde dicha década se han monitorizado las perdices de un área de estudio de 40 km2 en el condado de Norfolk en colaboración con la GWCT (Fundación para la Conservación de la Caza y la Fauna). A partir del año 2001, terratenientes, granjeros y guardabosques emprendieron un programa de creación de hábitat, suplementación alimentaria y control de los predadores, con el fin de recuperar las perdices. Las densidades aumentaron desde 4,7 parejas/km2 en marzo del 2001 a 54 parejas/km2 en marzo del 2011. Dichas densidades son comparables a las que había antes de la disminución nacional de los efectivos de perdiz pardilla. Durante los últimos tres inviernos se abatieron entre 13 y 74 aves/km2, y los efectivos primaverales continúan creciendo. Palabras clave: Perdiz pardilla, Creación de hábitat, Alimentación suplementaria, Control de predadores, Coto de caza. Received: 19 XII 11; Conditional acceptance: 27 II 12; Final acceptance: 25 V 12 R. A. H. Draycott, Game & Wildlife Conservation Trust, Fordingbridge, Hampshire, SP6 1EF, UK. E–mail: rdraycott@gwct.org.uk

ISSN: 1578–665X

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Introduction Historically, eastern England has been a stronghold for grey partridges Perdix perdix. From 1900 to 1920 between 20 and 72 birds/km2 were shot annually on shooting estates in eastern England (Tapper, 1992). However, between 1950 and 1990 shooting bags declined by 80% owing to a long–term dramatic decline in their population and range (Aebischer & Ewald, 2004). Between 1995 and 2009 populations continued to decline with estimates of a 54% decrease across the UK (47% in eastern England) over this period (Risely et al., 2011). There have been similar declines across the rest of its natural range and it is a species of European concern (PECBMS, 2010). The causes of the decline have been well researched and are mainly related to agricultural intensification and predation (Potts, 1986). More recently, it has been shown that habitat improvement and predation control can lead to increasing numbers of grey partridges at the local level (Aebischer & Ewald, 2010). In 1995, the UK Government designated grey partridge as a priority species under its Biodiversity Action Plan (BAP). It defined three targets for recovery: 1) to halt the decline by 2005; 2) to ensure the population is above 150,000 pairs by 2010; and 3) to enhance the current range. The aim of this paper is to describe the work undertaken by a large, privately–owned arable farming estate to restore wild grey partridges in an area where they were previously abundant. Methods Study area The restoration project consisted of approximately 40 km2 of farmland in the county of Norfolk in England. The study area is owned by one landowner, but includes some areas farmed by the landowner and other areas farmed by tenants. The study area was divided into five different game management units (beats), with gamebird management undertaken by one gamekeeper on each beat. The landscape is dominated by arable crop production, comprising winter and spring sown cereals (mainly wheat and barley), sugarbeet, oil–seed rape, peas (vining and combinable), potatoes and parsnips. These crops are typical of the region. Most fields are surrounded by hedgerows and grassy hedgebanks, (comprising approximately 2.5% of the land area) which are the favoured nest site for grey partridges in arable landscapes (Rands, 1986). There is a network of small woodlands, comprising 4.5% of the study area. Gamekeepers are employed on the estate to undertake predation control, habitat management and supplementary feeding. In addition to grey partridges, the study area is also managed to encourage wild pheasants (Phasianus colchicus), red–legged partridges (Alectoris rufa) and brown hare (Lepus europaeus). Habitat management The recovery project is based on three key requirements of grey partridges: 1) provision of suitable habitat

Draycott

for all aspects of the life history of the grey partridge; 2) protection from nest predators; and 3) provision of supplementary feed in winter and spring. Before set– side was abolished, it was utilised to provide habitat for partridges. Today, partridge habitat is provided via a combination of five or ten year environmental stewardship agreements, funded under the EU Common Agricultural Policy. Additional areas of habitat are privately funded by the landowner. These include grass margins (comprising 1.4% of the land area) to provide nesting cover, wild bird seed covers to provide winter cover and food and insect rich brood–rearing cover (comprising 3% of the land area). Wild seed mixtures are based on annual or biennial mixtures including cereals and brassicas. Brood rearing cover is sown close to nesting areas and consists of low input, spring sown cereal strips or perennial mixtures including chicory (Chicorium sp.) and lucerne (Medicago sativa). The structure of brood rearing cover is very important. Grey partridges will only use brood rearing areas if they feel safe, and if they can move freely through the vegetation. Therefore brood rearing cover should provide both an overhead canopy for protection from predators and an open structure at the base to allow freedom of movement (Sotherton & Swan, 2001). Predation control Common predators such as foxes (Vulpes vulpes), crow (Corvus corone), magpie (Pica pica) stoats (Mustela erminea) and rats (Rattus norvegicus) were controlled to reduce predation on adults, nests and broods in order to improve breeding success and population density of gamebirds (Tapper et al., 1996). Only predators which can be legally controlled were targeted, and gamekeepers adhered to all legal requirements, guidelines and codes of practice. Fox control consisted of night–time shooting and snaring, and corvids were shot or trapped using live catch Larsen traps. A network of tunnel traps were used to control small mammalian nest predators (e.g. brown rats and stoats). Supplementary feeding Although it has not been scientifically proven that spring supplementary feeding benefits grey partridges, there is scientific evidence that it benefits pheasants on farmland in Britain (Draycott et al., 1998; Draycott et al, 2005). Draycott et al. (1998) showed that it leads to improved body condition in nesting females, and Draycott et al. (2005) documented increases in breeding densities and improved recruitment in the autumn. In particular, hen pheasants with access to supplementary grain were much more likely to re–nest than unfed hens if their first nest was unsuccessful. Also, Hoodless et al. (2001) showed that pheasants provided supplementary grain spent much less time actively foraging for food than unfed hens. This could confer survival benefits for pheasants and partridges as less time spent feeding implies more time being vigilant. Supplementary feeding is provided for grey partridge from October or November (start date is dependent on environmental conditions) until the end of May. Feed hoppers are lo-

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1 km

Fig. 1. Location of breeding pairs of grey partridges (grey dots) on 40 km2 study area in Norfolk, England in 2000, prior to the recovery project. Fig. 1. Localización de las parejas de cría de perdiz pardilla (puntos grises) en un área de estudio de 40 km2 en Norfolk, Inglaterra, en el 2000, antes del proyecto de recuperación.

1 km

Fig. 2. Location of breeding pairs of grey partridges (grey dots) on 40 km2 study area in Norfolk, England in 2011. Fig. 2. Localización de las parejas de cría de perdiz pardilla (puntos grises) en un área de estudio de 40 km2 en Norfolk, Inglaterra, en el 2011.

384

Draycott

100 80 60 40 20 0 1956 1958 1960 1962 1964 1966 1968 1970 1972 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 2002 2004 2006 2008 2010

Grey partridge pairs/km2

120

Fig. 3. The mean ± SE densities of breeding pairs of grey partridge on an estate in Norfolk, England 1956–2011. Fig. 3. Densidades medias ± EE de las parejas de cría de perdiz pardilla en un coto de caza del estado de Norfolk, Inglaterra, 1956–2011.

cated approximately 75 m apart along hedgerows and beetle banks; the aim being to provide one hopper for every grey partridge territory. Draycott & Palmer (2008) showed that grey partridges tend to set up territories along hedgerows close to feed hoppers and that there was a positive relationship between pair density and the amount of hedgerow in the landscape.

Monitoring Population counts have been undertaken on the study site by the gamekeepers in conjunction with the GWCT since the 1950s. In March, all fields are surveyed with binoculars using a 4WD vehicle during the early morning or the evening (Potts, 1986). The location of

300

Partridges/km2

250

Partridges shot/km2 Partridges/km2

200 150 100 50

2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011

Fig. 4. The mean ± SE densities of partridges in autumn (solid line) and densities shot (dashed line) on an estate in Norfolk, England 2001–2011. Fig. 4. Densidades medias ± EE de perdices en otoño (línea continua) y densidades cazadas (línea de puntos) en un coto de caza del estado de Norfolk, Inglaterra, 2001–2011.

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Chick survival rate (%)

50 40 30 20 10 0

2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011

Fig. 5. The mean (± SE) chick survival rate of grey partridges on an an estate in Norfolk, England 2001–2011. Note: data missing for 2004. Fig. 5. Tasa de media (± EE) de supervivencia de crías de perdiz pardilla en un coto de caza del estado de Norfolk, Inglaterra, 2001–2011. No existen datos del año 2004.

observed breeding pairs and single birds is marked on a map. In recent years these have subsequently been entered into a GIS software package (figs. 1, 2). In the autumn, after harvest of most arable crops, all fields are re–surveyed. All coveys are counted, and the age of individuals (adult or juvenile) and the sex of the adult birds is recorded. All counts since 1999 were undertaken by the author, thereby removing any possibility of observer bias over the duration of the project (2001–2011). Each partridge beat (n = 5) was counted separately and numbers presented are mean values ± 1 SE. Results Response of grey partridges Numbers of grey partridges have increased dramatically during the course of the restoration period from (mean ± se) 4.7 ± 1.3 pairs/km2 in March 2001 to 54.2 ± 9.1 pairs/km2 in March 2011 (fig. 3). The estimated current average spring pair density on farms in Norfolk where intensive partridge management is not undertaken is 3.5 pairs/km2 (N. Kingdon, pers. com). In the highest density beat there were 88 pairs/ km2 in March 2011. In autumn, densities of greys have increased from 21.4 ± 6.2 birds/km2 in 2001 to 218.8 ± 26.8 birds/km2 in 2011 (fig. 4). The autumn figures are probably underestimates as approximately 25% of the study area cannot be counted in autumn due to crops (e.g. sugarbeet) still being present in the fields in September. However, figures are not adjusted to account for this as it is not known if partridges use this habitat in proportion to its availability.

Between 2001 and 2011 mean chick survival rate (calculated according to Aebischer & Reitz, 2000) was 33.7 ± 2.6 (fig. 5). Mean brood production rate (Aebischer & Reitz, 2000) between 2001 and 2011 was 89.9 ± 3.0%. In 2011, 74 partridges/km2 were harvested in the autumn. Discussion Under intensive, modern arable farming systems with no provision of brood rearing cover, chick survival rate is typically close to 20% (Aebischer & Ewald, 2004). Average chick survival rate over the course of the study was 33%, indicating that the provision of insect foraging areas has likely had a positive effect on chick survival rate. Mean brood production rate between 2001 and 2011 was 89.9 ± 3.0%, indicating low rates of nest predation owing to effective control of nest predators (Potts & Aebischer, 1991). In contrast to the rapid increase in numbers on this estate where intensive grey partridge management has been undertaken, on farms in England where no specific grey partridge conservation work is undertaken, grey partridges declined by 30% between 1999 and 2009 (Renwick et al., 2012). Other farms and estates that have undertaken management to restore partridges have also recorded increases in grey partridge numbers. For example, the GWCT Partridge Recovery Project at Royston in eastern England recorded a six–fold increase in numbers within five years, from 3 pairs/km2 in 2002 to 18 pairs/km2 in 2007 (Aebischer & Ewald 2010), and contributors to the GWCT Partridge Count Scheme have, on average, doubled the numbers of pairs

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counted between 2001 and 2010. (Aebischer & Ewald, 2010). Between 2008 and 2010 a maximum of 30% of the autumn stock was harvested on the study area (fig. 4). This is clearly within sustainable limits as the spring breeding stock continues to increase. The current density of grey partridges is the highest recorded with a modern commercial farming system (post–agricultural intensification) in the UK. It is therefore difficult to predict the optimum sustainable yield or, indeed, the carrying capacity of the land. However, Potts & Aebischer (1991) predicted through modelling an equilibrium density of 64 breeding pairs/km2 when nesting cover and chick food were not limiting factors and when nest predation rates were low. These results highlight the important role of private land managers in effective conservation of a declining species at the local level. The challenge is to translate these successes into partridge recovery at regional and national scales. References Aebischer, N. J. & Ewald, J. A., 2004. Managing the UK Grey Partridge Perdix perdix recovery: population change, reproduction, habitat and shooting. Ibis, 146 (Supplement 2): 181–191. – 2010. Grey Partridge Perdix perdix in the UK: recovery status, set–aside and shooting. Ibis, 152: 530–542. Aebischer, N. J. & Reitz, F., 2000. Estimating brood production and chick survival rates of grey partridges: an evaluation. Hungarian Small Game Bulletin, 5: 191–210. Draycott, R. A. H., Hoodless, A. N., Ludiman, M. N. & Robertson, P. A., 1998. Effects of spring feeding on body condition of captive–reared ring necked pheasants in Great Britain. Journal of Wildlife Management, 62: 557–563. Draycott, R. A. H., Woodburn, M. I. A., Carroll, J. P. & Sage, R. B., 2005. Effects of spring supplementary feeding on population density and breeding success of released pheasants Phasianus colchicus in Britain. Wildlife Biology, 11: 177–182. Draycott, R. & Palmer, J., 2008. Grey partridges and

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land use in Norfolk. Game & Wildlife Conservation Trust Review of 2007, 39: 28–29. Hoodless, A. N., Draycott, R. A. H., Ludiman, M. N. & Robertson, P. A., 2001. Spring foraging behaviour and diet of released pheasants (Phasianus colchicus) in the United Kingdom. In: Proceedings of the Perdix VII International Symposium on Partridges, Quails and Pheasants; Game and Wildlife Science, 18: 375–386 (M. G. Birkan, L. M. Smith, N. J. Aebischer, F. J. Purroy & P. A. Robertson, Eds.). Office National de la Chasse, Paris. PECBMS, 2010. Trends of common birds in Europe, 2010 update. European Bird Census Council, Prague. (www.ebcc.info/index.php?ID=387) Potts, G. R., 1986. The Partridge: Pesticides, Predation and Conservation. Collins, London. Potts, G. R. & Aebischer, N. J., 1991. Modelling the population dynamics of the Grey partridge: conservation and management In: Bird Population Studies: Relevance to Conservation and Management: 373–390 (C. M. Perrins, J. D. Lebreton & G. J. M. Hirons, Eds.). Oxford Univ. Press, Oxford. Rands, M. R. W., 1986. Effect of hedgerow characteristics on partridge breeding densities. Journal of Applied Ecology, 23: 479–487 Renwick, A. R., Eglington, S. M., Joys, A. C., Noble, D. G., Barimore, C., Conway, G. J., Downie, I. S., Risely, K. & Robinson, R. A., 2012. BirdTrends 2011. BTO Research Report 609. Risely, K., Renwick, A. R., Dadam, D., Eaton, M. A., Johnston, A., Baillie, S. R., Musgrove, A. J. & Noble, D. G., 2011. The breeding bird survey 2010. BTO Research Report 597. British Trust for Ornithology, Thetford. Sotherton, N. & Swan, M., 2001. Cover your broods. The Game Conservancy Trust Review of 2000, 32: 90–92. Tapper, S. C., 1992. Game Heritage: An Ecological Review from Shooting and Gamekeeping Records. Game Conservancy Ltd, Fordingbridge. Tapper, S. C., Potts, G. R. & Brockless, M. H., 1996. The effect of an experimental reduction in predation pressure on the breeding success and population density of grey partridges Perdix perdix. Journal of Applied Ecology, 33: 965–978.

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Every partridge counts, successful techniques used in the captive conservation breeding programme for wild grey partridge in Ireland K. Buckley, P. Kelly, B. Kavanagh, E. C. O’Gorman, T. Carnus & B. J. McMahon Buckley, K., Kelly, P., Kavanagh, B., O’Gorman, E. C., Carnus, T. & McMahon, B. J., 2012. Every partridge counts, successful techniques used in the captive conservation breeding programme for wild grey partridge in Ireland. Animal Biodiversity and Conservation, 35.2: 387–393. Abstract Every partridge counts, successful techniques used in the captive conservation breeding programme for wild grey partridge in Ireland.— Between 1998 and 2001 the last remaining wild grey partridge (Perdix perdix) population in Ireland faced imminent extinction with an estimated spring population of 4–6 pairs, and an autumn population of 22–24 birds. A captive breeding programme began in 2002 with two pairs of grey partridge. In the most successful year in 2010, 39 pairs produced a total of 510 chicks. Average chick survival rate was 65.13%. At 88.9 the highest chick survival rate was achieved in 2011. Chick survival of parent–reared birds in captivity is defined by the number of juveniles surviving at age six weeks: similar to estimations used for wild populations of grey partridge. Family coveys were released in late summer to early autumn. In most instances the entire family cohort was released as one unit. However, in coveys of twenty or above, an average of five parent–reared poults were held back as breeding stock for the following year. In early spring of the following year, birds held back were paired with single males or females trapped from the wild. The techniques we used were traditional and labour intensive but highly effective. We recommend that other grey partridge recovery projects should consider captive breeding using the methods employed in this programme to compliment other game management methods used. Key words: Grey partridge, Perdix perdix, Conservation breeding, Parent–rearing, Breeding success, Chick survival rate, Re–introduction. Resumen Cada perdiz cuenta, técnicas utilizadas con éxito en el programa de conservación de cría en cautividad para la perdiz pardilla en Irlanda.— Entre los años 1998 y 2001, los últimos restos de la población salvaje de perdiz pardilla (Perdix perdix) de Irlanda se enfrentaban a una extinción inminente, con una población primaveral estimada de 4–6 parejas, y una población otoñal de 22–24 aves. En el 2002 se inició un programa de cría en cautividad con dos parejas de perdices pardillas. En el año con mayor éxito, 2010, 39 parejas produjeron un total de 510 pollos. La tasa promedio de supervivencia de los pollos era del 65,13%. Se consiguió la mayor tasa de supervivencia de éstos en el 2011, que era del 88,9%. La supervivencia de las crías de los pollos de parejas de progenitores criados en cautividad se define mediante el número de jóvenes que sobreviven hasta la edad de seis semanas: parecida a las estimas utilizadas para las poblaciones salvajes de perdiz pardilla. Se soltaron grupos familiares desde finales del verano a principios del otoño. En la mayoría de los casos se soltaba la cohorte familiar entera como una unidad. Sin embargo, en los grupos de veinte o más, se retenía un promedio de cinco pollos criados por sus padres para formar la población de cría para el año siguiente. Al iniciarse la primavera del año siguiente, las aves retenidas se emparejaban con machos o hembras sueltos que se recogían de la naturaleza mediante trampas. Las técnicas que utilizamos eran las tradicionales y el trabajo intensivo, pero muy efectivo. Recomendamos que otros proyectos de recuperación de la perdiz pardilla consideren la cría en cautividad, utilizando los métodos empleados en este programa, para completar otros métodos de gestión utilizados. Palabras clave: Perdiz pardilla, Perdix perdix, Cría de conservación, Criado por los padres, Éxito reproductivo, Tasa de supervivencia de pollos, Reintroducción. ISSN: 1578–665X

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Received: 23 XII 11; Conditional acceptance: 30 IV 12; Final acceptance: 11 VI 12 Kieran Buckley, National Parks and Wildlife Service, Dept. of Arts Heritage & Gaeltach, Government Buildings, Athlumley, Kilcarin, Navan, County Meath, Ireland.– Paddy Kelly, Irish Grey Partridge Conservation Trust, Leamore Blueball County Offaly, Ireland.– Brendan P Kavanagh, Royal College of Surgeons in Ireland, 123 St. Stephens Green, Dublin 2, Ireland.– Edward Conor O’Gorman, British Association for Shooting and Conservation, Marford Mill, Rossett, Wrexham LL12 OHL, UK.– Barry J. McMahon & Tim Carnus, UCD School of Agriculture & Food Science, University College Dublin, Belfield, Dublin 4, Ireland. Corresponding author: K. Buckley. E–mail: peridix@hotmail.com

Animal Biodiversity and Conservation 35.2 (2012)

Introduction Re–introduction projects attempt to re–establish species within their historical ranges through the release of translocated wild or captive–bred individuals following extirpation or extinction in the wild (IUCN, 1998). Captive breeding of wild animals can be a vital component of re–introduction biology (Seddon et al., 2007). There are, however, a number of concerns regarding the potential success of captive breeding projects such as the reduced ability of individuals to survive in the wild (Snyder et al., 1996), in addition to poor health and condition of captive bred stock (Mathews et al., 2005). Captive breeding projects using individual birds from wild populations are not always guaranteed to succeed, e.g. Laysan Teal Anas laysanensis, (BirdLife International, 2010), and reservations have been expressed regarding the success of parent–rearing methods (Kreger et al., 2005). There is little published information relating to the captive breeding of wild grey partridge (Perdix perdix). The only reference on captive breeding was found in Maxwell (1911). Thus the information regarding captive breeding within Maxwell’s book formed the basis of the conservation breeding programme for the species in Ireland. The grey partridge is a farmland species which has declined across Europe (Burfield & Van Bommel, 2004; Kuijper et al., 2009). It is also a red–listed species in the Republic of Ireland and it is the most westerly population of the species in Europe (Lynas et al., 2007). The population has been in steep decline in Ireland since the middle of the 19th century (Ussher & Warren, 1900; Whilde, 1993). Although grey partridge were a popular game species in Ireland, anecdotal evidence suggests that few estates were managed specifically for grey partridge shooting. With little or no motivation to manage for shooting, fluctuations in the Irish population were inextricably linked to changes in agricultural land management. A population increase was noted from 1933 onwards,perhaps due to successful translocations although no reason is given and grey partridges had colonised areas in the west of Ireland where 20 years before they were unknown (O’Gorman, 2007). However, from the 1950s onwards across most parts of Ireland the grey partridge was still sparsely distributed, with most birds present in County Carlow (Kennedy et al., 1954). The status in the 1960s was similar to that given previously with no indication as to the actual size of the population. Ruttledge (1966) noted that the grey partridge was sparsely distributed and sometimes found in small cultivated fields of desolate agricultural areas (O’Gorman, 2001). Whilde (1993) described the native grey partridge in Ireland as ‘endangered’ with less than 200 breeding pairs. This estimation, however, should be considered with some caution as grey partridges were released from game farms in several areas across the country. In areas where no releases had taken place, local extinctions had already occurred decades earlier. For example, a number of release programmes were initiated on farmland in Kildare in 1993 and Wexford

389

in 1994, but neither resulted in the establishment of a successful wild breeding population (Kavanagh, 1998). In 1991, an autumn survey was carried out at the two last known wild partridge populations in the midlands of Ireland, at Boora, County Offaly & Lullymore and Kildare, both located in post–industrial cutaway bogs. An autumn survey of the two partridge populations was carried out at Boora, County Offaly and Lullymore and Ccounty Kildare. Both locations are cutaway peat land bogs. Cutaway bog is an open, mostly barren habitat following industrial peat extraction. Re–colonisation by a variety of plant communities formerly found in traditional tillage fields emerged in subsequent years. From 1992 onwards, a combination of surveys and information on sightings was collated to produce minimum estimates of the total autumn population for these two sites (Kavanagh, 1992, 1998). From 1996 to 2002, no releases of grey partridge had occurred in Boora. This population was monitored by spring and autumn estimations only. In 1993, three grey partridge pairs were radio–tagged in Lullymore (Hearshaw, 1996), initiating the first focused study of the species in Ireland. Nesting sites were chosen in areas of recolonized cutaway bogs, but all were predated (O’Gorman, 2007). In spring 1994, the Lullymore population was estimated at only 3–4 pairs, and by 2000, grey partridge were extinct in Lullymore. Thus, after the year 2000, the only wild population left in the Republic of Ireland was at Boora. In 1996 a conservation project, including predator control and habitat management with concurrent research, was established to prevent the extinction of this very last Irish population. With an increase in the total autumn population from 59 in 1996 to 72 birds in 1997, the initial response of the grey partridge recovery project was encouraging. However, three successive wet and cold summers followed, and together with a further loss of habitat owing to reclamation of the cut–away for grassland and forestry by Bord Na Mona —an Irish semi–state company set up in the mid 1950’s to exploit Irish peat lands for fuel— the Boora population declined to 22 birds in autumn 2001. The extinction of the species as a breeding bird in Ireland therefore seemed imminent. Since the re–introduction of captive–bred artificially–reared birds has been shown to be ineffective due to high mortality (Potts, 1986; Rantanen et al., 2010; Buner et al., 2011), the release of game–farm reared grey partridge was not considered. Instead, a behavioural study on the ability of captive grey partridges was carried out in Boora (Kavanagh & Fattbert, 2002) using game–farm reared breeding stock for parent–rearing. This research concluded that the average clutch size of artificially–reared birds (35.5) was almost double the average clutch size of wild birds and hatching success of game farm–reared birds was virtually zero. Any hatched chicks were compromised by the degenerate and maladapted behaviour of their parent birds. As a result of the difficulties associated with captive–bred artificially–reared birds and their subsequent lack of success in the wild (Rantanen et al., 2010; Buner & Schaub, 2008; Buner et al., 2011), a decision

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was taken to proceed with a conservation breeding programme using wild–caught grey partridges from Estonia and Boora (see also 'Origin of founder captive stock for parent–rearing' in Material and methods). This paper reflects on the successful methods used to breed wild grey partridge in captivity in Ireland and the relevance for reintroduction and restocking of the species for conservation purposes. In addition, some recommendations for the successful breeding of wild grey partridges in captivity are outlined. Material and methods Study area The captive breeding experiment was located in the area of the grey partridge conservation project in Boora, County Offaly, Ireland. The study comprised a 5 km2 core area. However, a variable number of grey partridge were dispersing 12 km2 from the core area, particularly to the east and south where the tillage and root crop farming was practised. Outside the core area, non–intensive predator control was also carried out, namely the culling of localised fox populations. This core site was chosen on the basis of logistical prudence and its convenience to the local population of wild grey partridges. In terms of managing the experiment, the site at Boora provided a source of wild grey partridge for trapping. Simultaneously, predators were controlled systematically and suitable habitats were created and managed for grey partridge post–release. Rearing methods Origin of founder captive stock for parent–rearing During the first three years, from 2002–2004, grey partridges used in our captive parent–reared breeding programme were obtained from a similar captive programme in France (Kavanagh, 2001). These birds were the progeny of wild trapped birds. They had not been subjected to successive generations in captivity, unlike those of normal game farm stock. In 2005, grey partridges used in our captive breeding programme were trapped directly from the wild in Estonia. The population in Ireland shares ancestry with both eastern and western populations of grey partridges in Europe (Liukkonen–Anttila et al., 2002). These birds were paired in captivity with birds from the local Irish population (see below). Estonian birds comprising pure pairs (both individuals from Estonia) or mixed pairs (one local Irish and the other Estonian) proved far more difficult to manage. Thus, in 2005, when we used either pure or mixed pairs, we recorded the lowest chick survival rate. Determined efforts were made to understand why chick survival was so low. After careful and prolonged observations it was discovered that Estonian adult birds and/or mixed pairs spent a significant amount of time hidden in cover, refusing to bring their chicks to the high protein chick crumb provided. Any journeys that were made by adult birds with their chicks were short lived. Thus an

Buckley et al.

insufficient amount of high protein food was consumed by the chicks. Consequently, chick mortality increased significantly. Wild–caught adult birds from Boora for parent–rearing Each year grey partridges from the project area in Boora were trapped from the wild. Trapping did not begin until after wild birds had paired. The reason for this approach was not to interfere with the mating behaviour of the few remaining last wild birds. Trapping for the captive breeding programme focused mainly on un–paired males; single females were only occasionally caught as they normally re–pair very quickly in the wild owing to a male surplus that is typical in wild partridge populations. Single birds were attracted to a pen containing a live decoy of the opposite sex, which was chosen from a number of birds which were retained in captivity from the previous year in a small 1 m2 container within a larger pen in which the trapped bird was visible. The door of the pen was left open to allow the wild bird to enter, triggering the door to shut behind. Trapped birds were placed in a pentagon shaped pen, each section measuring 2.5 m long by 1.2 m high. These pens were situated on a free–draining grassland and post–glacial bolder clay. Depending on the circumstances, either three domesticated males or females, which were kept in captivity over–winter, were then introduced to the wild trapped bird to allow for unforced pairing. Mate choice was noted by the keeper based on the behaviour observed, e.g. a male and female staying together or the newly formed pair driving other un–paired birds away from them. Each breeding pair was then given an individual pen, similar to the one described above. The normal breeding process began in captivity, e.g. nest construction, egg–laying and incubation. After hatching, chicks were allowed 24 hours to digest their natural egg yolk food supply. Following this period, chicks were caught up, followed by their parent birds, in that order. The family unit was then placed in a brood–rearing box measuring 1.5 m long by 0.6 m high by 0.6 m wide. The box was covered with a dust proof green mesh to prevent injury. A clear sheet of acrylic glass was placed at one end of the box over the top one–third of the unit. The family group was kept in this box for a period of one week to ten days; grey partridge chick starter crumb and water was provided at all times. It was decided to bring the crumb to the chicks instead of attempting to get the chicks to come to the crumb after the learning experience with the Estonian birds described above. Chicks confined in these boxes were effectively trained to the crumb. The pair and their chicks were checked regularly for short periods of time throughout each day of their confinement. Contact with chicks was kept to an absolute minimum to prevent habituation. Family groups held for longer than one week were first fed on a diet of grey partridge chick starter crumb, followed by mini–pellets and ‘finished off’ on grower pellets prior to release as an intact family covey. After time periods ranging from one to six weeks, the leg ringed adult pair was released with their chicks into suitable habitat.

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Table 1. Captive breeding data from 2002 to 2011 of the Irish grey partridge recovery programme in Boora. Tabla 1. Datos de cría en cautividad del 2002 al 2011, en el programa de recuperación de perdiz pardilla en Boora. Year 2002

2003

2004

2005

2006

2007

2008

2009

2010

2011

–

116

127

–

187

281

510

307

Number of pairs nesting 2

Number of pairs incubating eggs 2

Number of total chicks hatched –

Number of chicks surviving to 5 weeks 10

–

139

185

436

273

76.7

10.3

59.1

–

74.3

65.8

85.5

88.9

Chick survival rate (%) –

60.6

Rearing with bantam chickens

Disease and disease prevention of rearing stock

When wild grey partridge pairs were confined to a breeding pen, the majority of their eggs were laid in a scrape. These eggs were covered by the pair with dead vegetation, until incubation commenced. However, our experience indicated that females were not inclined to lay all their eggs in one nest. Eggs not laid in the nest were usually 'scattered' around the pen. Left unused these eggs represented an unacceptable waste of potential adult birds. Thus, these eggs were 'bled out' until a suitable number was obtained. When approximately 20 eggs were collected they were placed under a mongrel broody Bantam chicken for incubation. Standard facilities were provided to the bantams to rear these chicks until they were five weeks old. Juvenile bantam–reared grey partridges were then isolated from their foster mother for a period of one week. These juveniles were leg–ringed and placed in a fostering unit. The fostering unit was then moved to an area where a single adult bird or a barren breeding pair was present in the wild. These pairs or single birds quickly fostered the juvenile bantam–reared grey partridges as their own as described in detail in Buner & Aebischer (2008). In addition to fostering juvenile birds, grey partridge coveys were removed from their breeding pens and placed into a square 1.5 m x 1.5 m2 wooden box with a hinged front door. The coveys were then transported to release sites with suitable habitats. A 20 m string was then attached to the door. After a period of 30 minutes the hinged door was opened, allowing the covey to escape un–stressed.

Throughout the entire period of the captive breeding programme, the grey partridges remained susceptible to disease. Diseases included coccidiosis, septicaemia, lung consolidation causing Escherichia coli septicaemia, and gapes (Syngamus trachea). These diseases were identified by a veterinary surgeon in post mortems on 28–day old chicks and one adult (Clerkin, 2008). Over the period of the captive breeding programme some mortality occurred in chicks less than one week old. On veterinary advice, a four–day treatment of vitamins and anti–pathogenic medicines including Amoxinsol–50 and water soluble vitamin E was administered. In addition, anti–gape treatment was given to chicks by mixing flubenvet into chick starter crumb, mini pellets and grower pellets. This approach significantly reduced the mortality of chicks and juveniles. To further reduce the incidence of pathogenic and parasitic disease, breeding pens were moved each year onto ‘fresh ground’ prior to the onset of the breeding season. Results Captive breeding data from 2002 to 2011 are summarised in table 1. There was a substantial increase in chick survival rate (CSR) over time. A CSR of 88.9% in 2011 represents the highest chick survival rate ever recorded for the captive breeding programme. The lowest CSR recorded was 10.3% in 2005 which appears to be an outlier as over the 10–year period owing to the reasons

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explained under 2.2., the average CSR was 65.2%. No data were recorded in 2007 due to a temporary shortage of staff. Discussion Captive breeding of wild animals may be a vital component of re–introduction biology but success is not a foregone conclusion (Snyder et al., 1996; Mathews et al., 2005). The ability of wild individuals to survive in a captive environment is a major concern. The balance of maintaining good health and condition in conjunction with fostering a sufficient reproductive breeding success to reintroduce individuals cannot be overstated. In addition, there is an underlying necessity to maintain the innate wild behaviours and instincts of the breeding stock. Where population augmentation is the aim, integration of the most appropriate genetic variation is another consideration.To our knowledge the Irish captive breeding strategy of wild grey partridges is the most successful captive breeding programme aimed at re–establishing a wild population ever recorded,with a CSR of 88.9% in 2011 and an average CSR of 65.2% over the duration of 10 years to date. The most plausible explanation for the success of this programme is the result of numerous factors employed by the personnel involved. However, the action taken to bring the crumb to the chicks and the attention paid to the prevention of disease is fundamental to the success of the breeding programme in Boora. It would also appear that taking into account the wild behaviour of the partridges in captivity and adjusting the breeding techniques to incorporate this behaviour is paramount to the success of the birds’ post–release survival and successful breeding. Captive breeding widens the scope for the re–introduction of endangered bird species such as the grey partridge in Ireland. This strategy ensures that insofar as possible, each individual bird within an endangered population can make a contribution to the population as a whole. Breeding in captivity in mixed pairings with wild and domesticated individuals can produce high numbers of well–adapted offspring which are ideal for the release into the wild. This is reflected in an overall increase in the wild population. Since the inception of the captive breeding programme the autumn populations of wild grey partridge has increased on average 42.8% over the ten–year period It can be assumed that our released parent–reared juveniles benefited directly in terms of survival from the local knowledge of their captured wild parent (either male or female in each parent–reared covey released) and that the subsequent breeding success of these offspring is making a major contribution to the wild population. Captive breeding also provides the opportunity to maximise the breeding productivity of wild grey partridges in years when the variables of an Irish summer or indeed the lack of continuity of funding reduce the wild productivity below sustainable levels. In Ireland the captive breeding programme has not only assisted the recovery of the species in Boora but the success has resulted in facilitating the relo-

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cation or re–introduction of coveys of grey partridge into additional farmland habitats formerly occupied by the species. The methods used in this breeding programme can be applied in other translocation, restocking or re–introduction projects throughout Europe in areas where the species used to be present or where critically endangered wild populations would benefit from population augmentation. However, in Ireland and indeed in other countries, captive breeding of wild grey partridges is not a long–term solution. Captive breeding and release of grey partridges should only be implemented in conjunction with habitat management and predator control. From a European and Irish perspective, the long–term viability of this iconic farmland specialist species is to create agricultural ecosystems which can meet its ecological requirements. It would appear that the most likely vehicle to achieve this is to devise targeted agri–environment schemes. Although the captive breeding programme has been successful, the long–term viability of the population of grey partridges, as with many farmland specialist birds in Ireland, is more likely to be dependent on agri–environment policy (McMahon, 2007). Captive breeding was used as a conservation tool in Ireland to augment the existing breeding population of grey partridge because without it, this enigmatic species would most likely now be extinct. Conservation implications The success of the breeding programme has enabled recommendations to be devised as a result of the lessons learned: (1) Captive–bred hand–reared grey partridge should never be used as they are behaviourally maladapted and do not breed successfully post release. (2) To maximise survival of released coveys, parent birds should contain at least one wild–caught grey partridge and ideally, the captured bird originates from the release site; as a rule, the caught wild birds bring their survival experiences to the released covey and hence increase post–release survival of the whole group. (3) It is recommended that a wild male is allowed to pair naturally with a domesticated female; 24 hours post hatching, broods of chicks should be placed into brood–rearing boxes with parent birds following in that order. (4) Ready access to crumb for the chicks up until they are at least ten days old is essential. (5) Contact between game keepers and the breeding pairs and their chicks should be kept to an absolute minimum to avoid habituation. (6) Wild grey partridges should not spend more than one year in captivity. (7) Good habitat, supplementary feeding and predation control should be in situ before any releases occur. Acknowledgements We would like to acknowledge the National Parks & Wildlife Service of Ireland who have funded the recovery of the species from the outset. Particular thanks to the North Eastern Region of NPWS. We would also like to acknowledge the crucial role played by Mr Val Swan.

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We extend our gratitude to Padraig Comerford, James Moore, Noel Bugler, Colm Malone, Ciara Flynn. Ciaran O Keeffe,, David Tierney and Jerry Lecky NPWS at Ely Place, Dublin 2. We would like to thank John Walsh of The Irish Grey Partridge Conservation Trust and the, Native Species Conservation Committee of Dublin & Belfast Zoo’s and Fota Wildlife Park. Thanks also to Des Crofton and Simon Deveraux of The National Association for the Regional Game Councils of Ireland, Andres Lillamae of The Estonian Hunters Society and the Game and Wildlife Conservation Trust UK. References Armitage, D., 2009. Research Report on the Irish Grey Partridge Population Prepared for the National Parks and Wildlife Services. Ely Place Dublin 2, Ireland. BirdLife International, 2010. Anas laysanensis. In: IUCN 2011. IUCN Red List of Threatened Species. Version 2011.2. <www.iucnredlist.org>. Downloaded on 13 December 2011. Buner, F. & Schaub, M., 2008. How do different releasing techniques affect the survival of reintroduced grey partridges Perdix perdix. Wildl. Biol., 14: 26–35. Buner, F. & Aebischer, N., 2008. Guidelines for re–establishing grey partridges through releasing. Game and Wildlife Conservation Trust, Fordingbridge. ISBN: 978–1–901369–17–5. Buner, F., Browne, S. & Aebischer, N., 2011. Experimental assessment of release methods for the re–establishment of a red–listed galliform, the grey partridge (Perdix perdix). Biological Conservation, 44: 593–601. Burfield, I. & Van Bommel, F., 2004. Birds in Europe: Population Estimates, Trends and Conservation Status. Birdlife International, Cambridge. Clerkin, F., 2008. Post mortem report on grey partridge to the National Grey Partridge Conservation Project, C/o Desmond Crofton, Ranelagh Dublin, 6. Hearshaw, J., 1996. The ecology of the grey partridge (Perdix perdix) in the Irish Midlands. M. Sc. Thesis, Univ. College Dublin, Dublin. IUCN (World Conservation Union), 1998. Guidelines for re–introductions. IUCN /SSC Re–introduction specialist group, IUCN, Gland, Switzerland, and Cambridge, United Kingdom. Kavanagh, B., 1992. Irish Grey Partridge (Perdix perdix) population survey 1991, with special reference to population and habitat use in cutaway bogland. Gibier Faune Sauvage, 9: 503–514. – 1998. Can the Irish grey partridge (Perdix perdix) be saved? A national conservation strategy. In: Perdix VII, International symposium on partridges, quails and pheasants, Gibier Faune Sauvage, October 1995, Dourdan, France, 15: 533–546 (M. Birkan, L. M. Smith, N. J. Aebischer, F. J. Purroy & P. A. Robertson, Eds.). – 2001. The 2000 Annual Report of the Grey partridge Conservation Project. Unpublished Report to National Parks and WildlifeService Ely Place Dublin 2 Ireland.

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Kavanagh, B. & Fattbert, K., 2002. Behavioral study of game farmed grey partridges ability to breed, unpublished report of the Irish Grey Partridge Conservation Trust, Cromwellstown Kilteel Naas County Kildare. Kennedy, P. G, Ruttledge, R. F. & Scroope, C. F., 1954. The Birds of Ireland. Oliver and Boyd Ltd., Edinburgh. Kuijper, D. P. J., Oosterveld, E. & Wymenga, E., 2009. Decline and potential recovery of the European grey partridge (Perdix perdix) population – a review. Eur. J. Wildlife Res., 55: 455–463. Kreger, M. D., Hatfield, J. S., Estevez, I., Gee, G. F. & Clugston, D. A., 2005. The effects of captive rearing on the behavior of newly–released Whooping Cranes (Grus americana). Appl. Anim. Behav. Sci., 93: 165–178. Liukkonen–Anttila, T., Uimaniemi, L., Orell, M. & Lumme J., 2002. Mitochondrial DNA variation and the phylogeography of the grey partridge (Perdix perdix) in Europe: from Pleistocene history to present day populations. J. Evolution. Biol., 15: 971–982. Lynas. P., Newton S. F. & Robinson, J. A., 2007. The status of birds in Ireland: an analysis of conservation concern 2008–2013. Irish Birds, 8: 149–166. Mathews, F., Orros, M., McLaren, G., Gelling, M. & Foster, R., 2005. Keeping fit on the ark: assessing the suitability of captive–bred animals for release. Biological Conservation, 121: 569–577. Maxwell, A., 1911. Partridges and Partridge Manors. London, Black. McMahon, B. J., 2007. Irish Agriculture and farmland birds: research to date and future priorities. Irish Birds, 8: 195–206. O’Gorman, E. C., 2001. Home range and habitat use of the endangered grey partridge (Perdix perdix) in the Irish midlands. Ph. D. Thesis, Trinity College, Dublin, Ireland. – 2007. Overview report to the National Grey Partridge Conservation Project: un–published report to the National Grey Partridge Conservation Company, Dublin 6, Ireland. Potts, G. R., 1986. The Partridge: Pesticides, Predation and Conservation. Collins, London. Rantanen, E. M., Buner, F., Riordan, P., Sotherton, N. & Macdonald, D. W., 2010. Habitat preferences and survival in wildlife reintroductions: an ecological trap in reintroduced grey partridges. J. Appl. Ecol., 47: 1357–1364. Ruttledge, R. J., 1966. Ireland’s birds. Witherdy Ltd., London. Seddon, P. J., Armstrong, D. P. & Maloney, R., 2007. Developing the Science of Reintroduction Biology. Conservation Biology, 21: 303–312. Snyder, N. F. R., Derrickson, S. R., Beissinger, S. R., Wiley, J. W., Smith, S. B., Toone, W. D. & Miller, B., 1996. Limitations of captive breeding in endangered species recovery. Conserv. Biol., 10: 338–348. Ussher, R. J. & Warren, R., 1900. The birds of Ireland. Gurney and Jackson, London. Whilde, A., 1993. Threatened Mammals, Birds, Amphibians and Fish in Ireland. Irish Red Data Book 2: Vertebrates. H.M.S.O.

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Does fox control improve red–legged partridge (Alectoris rufa) survival? An experimental study in Northern Spain A. Mateo–Moriones, R. Villafuerte & P. Ferreras

Mateo–Moriones, A., Villafuerte, R. & Ferreras, P., 2012. Does fox control improve red–legged partridge (Alectoris rufa) survival? An experimental study in Northern Spain. Animal Biodiversity and Conservation, 35.2: 395–404. Abstract Does fox control improve red–legged partridge (Alectoris rufa) survival? An experimental study in Northern Spain.— This work evaluates the effectiveness of fox control as a method to improve the survival of red–legged partridge (Alectoris rufa). We radio–tracked 89 adult partridges and their chicks (62 few days old chicks and 46 over one–month–old chicks) and monitored their nests (N = 45) on two hunting estates in northern Spain over two years. Generalist predators (red fox, Vulpes vulpes, and magpie, Pica pica) were selectively controlled on one half of each estate during the first year, and on the other half in the second year. We estimated the effect of predator control on survival rates. Predator control did not improve survival rates for adult partridges and nests, but it improved chick survival, especially for chicks over one–month old. Key words: Predator control, Radio–tracking, Red–legged partridge, Red fox, Survival rates. Resumen ¿Puede el control de zorros mejorar la supervivencia de la perdiz roja (Alectoris rufa)? Un estudio experimental en el Norte de España.— Evaluamos la efectividad del control selectivo de zorros como método para mejorar la supervivencia de la perdiz roja (Alectoris rufa). Para ello, radio–seguimos 89 perdices adultas y sus pollos (62 pollos de pocos días y 46 pollos de más de un mes de edad), e hicimos un seguimiento de sus nidos en dos cotos de caza del Norte de España durante dos años. En la mitad de la superficie de cada coto se controlaron de forma selectiva los depredadores generalistas (zorro, Vulpes vulpes, y urracas, Pica pica) durante el primer año, y los tratamientos se invirtieron entre zonas durante el segundo año. Estimamos el efecto del control de depredadores sobre las tasas de supervivencia. El control de depredadores no mejoró la supervivencia de los adultos y nidos de perdiz, pero sí mejoró la supervivencia de los pollos, especialmente para los pollos de más de un mes de edad. Palabras clave: Control de depredadores, Radio–seguimiento, Perdiz roja, Zorro común, Tasas de supervivencia. Received: 4 I 12; Conditional acceptance: 9 V 12; Final acceptance: 11 VI 12 Ainhoa Mateo–Moriones, Rafael Villafuerte & Pablo Ferreras, Inst. de Investigación en Recursos Cinegéticos (IREC, CSIC–UCLM–JCCM), Ronda de Toledo s/n., 13071 Ciudad Real, España (Spain). Corresponding author: Ainhoa Mateo–Moriones, E–mail: ainhoamm@yahoo.es

ISSN: 1578–665X

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Introduction The red–legged partridge (Alectoris rufa) is a galliform species distributed in southwestern Europe (Iberian Peninsula, France, Italy) and the UK. It is globally considered as 'vulnerable' (Aebischer & Potts, 1994) and as a 'species of special interest' at the European level (Tucker & Heath, 1994). During recent decades, red– legged partridge populations have declined notably (Cramp & Simmons, 1980; Aebischer & Potts, 1994; Blanco–Aguiar et al., 2003). The causes of decline in Spain are multiple, including habitat loss (Buenestado et al., 2008), pathogens (Millán et al., 2001; Villanúa et al., 2008; Díaz–Sanchez et al., 2011) and genetic introgression resulting from restocking with farm– reared partridges (Negro et al., 2001; Barilani et al., 2007; Blanco–Aguiar et al., 2008), excessive hunting pressure (Blanco–Aguiar et al., 2003) and predation (Moleón et al., 2008; Buenestado et al., 2009). Control of predator species can reduce predation suffered by game species and facilitate their recovery (Tapper et al., 1996; Smith et al., 2010). However poorly designed predator control could be counterproductive, since it could induce high densities of small predators by a process of mesopredator release (Crooks & Soulé, 1999; Blanco–Aguiar et al., 2001; Beja et al., 2009). An indiscriminate or unselective predator control may also affect other species and lead to an imbalance of natural ecosystems (Coté & Sutherland, 1997). Despite these concerns predator control is a widespread game management practice in Spain (Delibes– Mateos et al., 2009); its effectiveness on improving the demographic parameters of small game species is still unknown. In the case of the red–legged partridge, a few studies have focused on the effect of predator control on nest survival (Yanes et al., 1998; Herranz, 2000), but studies of the effect on survival of adult and chick survival are lacking. Our aim was to evaluate the effectiveness of predator control as a tool to improve the survival of several age classes of the red–legged partridge. We evaluated the effect of predator control on the survival of adults, nests and partridge chicks during a two–year experimental study on two hunting estates in Northern Spain. Material and methods Study areas This study was carried out over two consecutive years (2008 and 2009) on two hunting estates in the southwestern part of Navarra, Northern Spain: Arroniz (Study Area 1: 5,477 ha) and Sesma (Study Area 2: 7,067 ha), both with similar environmental and social characteristics. Most of the land area of these estates is covered by arable crops (> 70%), with natural vegetation consisting of Mediterranean scrubland with some pine plantations (See table 1 for further information). The main game species are red–legged partridges, European rabbits (Oryctolagus cuniculus) and Iberian hares (Lepus granatensis). Both study areas have medium abundances of red–legged partridge (spring

Table 1. Main landscape characteristics: habitat surface in hectares (ha) and % in parentheses. Tabla 1. Principales características del paisaje: superficie en hectáreas (ha) y % entre paréntesis. Landscape characteristics Study Area 1

Study Area 2

Surface covered (ha)

5,518.7

7,111.0

Mediterranean scrubland/forest (%) 1,050.4 (19%)

1,868.5 (26.3%)

Irrigated croplands (%)

22.9 (0.4%)

161.6 (2.3%)

Unirrigated croplands (%) 4,387.0 (79.5%)

5,013.0 (70.5%)

Total arable croplands (%) 4,409.9 (79.9%)

5,174.6 (72.8%)

Unproductive land/fallow

58.4 (1.1%)

67.8 (1.0%)

Diversity index (Shannon H)

0.248

0.316

Average patch size (ha)

17.7

23.16

Kilometer Abundance Index, KAI, number of individuals recorded per kilometre travelled: 1.5). Both study areas are extensive game estates (low to medium hunters density), where partridges are hunted by walk–up shooting with dogs. Game management is carried out by one gamekeeper in each estate and consists of water supply, and some areas of reserve/refuges where hunting is not practised. Farm–bred partridges were not released during the study period and no artificial feeding was provided. The most important predators for the red–legged partridge vary in abundance; red foxes (Vulpes vulpes) are in high abundance (summer KAI 0.14–0.22 in Study Area 1 and 0.25–0.28 in Study Area 2) and magpies (Pica pica) are in low abundance (spring KAI 0.06 in Study Area 1 and 0.16 in Study Area 2). Other potential partridge predators include diurnal raptor species, such as Montagu’s harrier (Circus pygargus), hen harrier (Circus cyaneus), marsh harrier (Circus aeruginosus), golden eagle (Aquila chrysaetos) and booted eagle (Hieraaetus pennatus), and nocturnal raptors such as eagle owl (Bubo bubo) and short–eared owl (Asio flammeus). Each study area was divided into two treatment zones: a predator control zone (hereafter PC) and non–predator control zone (hereafter NPC). During the first year of the study (2008), the gamekeepers

Animal Biodiversity and Conservation 35.2 (2012)

from each hunting estate selectively and intensively controlled red foxes in the PC zone, while no predator control was applied in the NPC zone. Treatments were reversed between zones during the second year of the study (2009). Predator control was performed from February to December by authorized staff (gamekeepers and some hunters from the hunting societies) only, using legal methods. During the hunting season, hunters were also authorised to shoot red foxes during their hunting activity. We planned to control magpies at the beginning of the study as they are considered one of the main predators of red–legged partridge nests and chicks (Yanes et al., 1998). A total of 86 magpies were culled during the study (60 in 2008 and 26 in 2009; table 2). However, the background magpie abundance in these areas is very low (maximum KAI recorded values: 0.06 in Study Area 1 and 0.16 in Study Area 2 in autumn 2008), making it difficult to assess the effect of controlling the abundance of this species.

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Table 2. Number of foxes (Fx) and magpies (Mg) harvested during each study year in the predator control zones of Study Area 1 and 2: *Game managers in Study Area 1 decided not to control magpies in 2009 due to their low abundance in the control zone that year. Tabla 2. Número de zorros (Fx) y urracas (Mg) extraídos durante cada año en las zonas de control de depredadores del área de estudio 1 y 2: *Durante 2009 los gestores del área de estudio 1 decidieron no realizar control de urracas debido a su baja abundancia en la zona de control de depredadores correspondiente a ese año.

2008

2009

Study Area

Monitoring predator populations

Fox populations were monitored using spotlight counts, performed once a month, from spring (March–April) to autumn (October). Magpie populations were monitored by diurnal counts, carried out twice a year (April and October). In both cases, KAIs were calculated.

Total

Capture and radio–tracking of adult and chick partridges and location of nests From February to the end of April, adult partridges were captured in all study areas, using two methods: (i) cages with a living decoy (Casas et al., 2009) and (ii) night captures using a net and a spotlight (Buenestado et al., 2009; Casas et al., 2009). Adult partridges were radio–tagged with a collar transmitter model TW–3 (11 g of weight, Biotrack Ltd, Dorset, UK) and radio–tracked every 24–48 hours, either until the transmitter batteries ran out (at around eight months) or at the end of the annual tracking period (November). The radio–tracking of adult partridges allowed us to locate their nests during the breeding season. Nests were geo–referenced and monitored from a distance using binoculars to scan the nest location, trying not to disturb the hen during the incubation period. Hatching dates were recorded if the nest succeeded, and cause and date of failure were noted if the nest was unsuccessful. Once hatched, chicks were captured at two different ages: (i) chicks between one and four days old (Chicks_1) and (ii) chicks over a month of life (Chicks_2). Those chicks captured earlier (Chicks_1) were radio–tagged with small radio–transmitters (model PIP21, 0.45 g weight, Biotrack Ltd., Dorset, UK) glued to their back (Mateo– Moriones et al., 2012) and located daily. Chicks_2 were radio–tagged with transmitters (model TW–41, 4.5 g. weight, Biotrack, Ltd., Dorset, UK) placed dorsally as a backpack (Mateo–Moriones et al., 2012) and located every 1–3 days. According to tests both in captivity and in the field, these tagging methods seem to have reduced effects on chicks’ survival (Mateo–Moriones

et al., 2012). Radio–tracking of partridges (adults and chicks) and monitoring of nests provided the data required for estimating survival rates and cause of mortality of adult partridges, nests and chicks. Estimation of survival rates of adult partridges, nests and chicks Survival rates were estimated for each age group by using the most appropriate application in program MARK 4.0, based on the available data (White & Burnham, 1999; Rotella et al., 2004). MARK provides estimates of survival rates and allows the comparison of different models with combinations of factors or variables to the observed survival data. The models are ranked according to their explanatory ability using the Akaike Information Criterion (AIC, White & Burnham, 1999). Adult survival was estimated with the 'Known Fate' application; candidate factors included in the models were the time of year (1. Winter coveys; 2. Mating; 3. Nesting; and 4. Broods), year (2008/2009), study area (1 and 2), treatment effects (PC. Predator control/ NPC. Non–predator control), sex (male/female), and age class (subadult/adult). Nest success was estimated with the 'Nest Survival' application, with model candidate factors: year, study area, treatment, sex of the adult incubating the nests (both males and females can incubate, Casas et al., 2009) and laying period (early/late, before, or after the median laying date). The survival of the chicks during the first two weeks of life (Chicks_1) was estimated with the 'Nest Survival' application (Moynahan et al., 2007), including as factors: year, study area, treatment, age and brood identity (in order to control for non–independence of

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NPC zone

0.6

PC zone

Fox KAI

0.5 0.4 0.3 0.2 0.1 0 March

May

July

August

September

October

0.6

Fox KAI

0.5 0.4 0.3 0.2 0.1

0 April

July

August

September

October

Fig. 1. Fox Kilometric Abundance Indexes (KAI) in the zones where predator control was applied (PC zone, continuous line with squares) and in the zones where it was not applied (NPC zone, discontinuous line with triangles) during 2009, in Study Area 1 (A) and Study Area 2 (B). Fig. 1. Índices kilométricos de abundancia de zorros (KAI) en zonas donde se aplicó control de depredadores (zona PC, línea continua con cuadrados) y en zonas donde no se aplicó (zona NPC, línea discontinua con triángulos) durante el año 2009 en el área de estudio 1 (A) y el área de estudio 2 (B).

survival of sibling chicks). For chicks over 1 month of life (Chicks_2) we used the 'Known Fate' application (Cooch & White, 2010), with the following factors in the models: year, study area, treatment and weight of the chicks at capture (as a correlate of their age). In each case, models were ranked according to the value of the AIC corrected for small samples (AICc). Those models with a ΔAIC < 2 with respect to the lowest AIC were considered as the most plausible models for explaining the observed data. Our main goal was to determine whether the model including the treatment (predator control) was included among the most plausible models. This would indicate that predator control affects the survival rate being considered in the model. Moreover, the overall weight of treatment as a factor explaining the survival rates was calculated considering the partial weights of all the models including such factor. We also estimated the survival rates of adult partridges, nests and chicks for each treatment.

There was some uncertainty concerning the fate of some of the chicks, as we lost the signal of some transmitters during the radio–tracking period. In order to take into account such uncertainty we considered two extreme scenarios: (i) the minimum survival scenario that assumes that all signal losses were due to the death of the chicks, and (ii) the maximum survival scenario that assumes that all the signal losses were due to transmitter failure or early exhaustion of batteries, not implying the chick death. We assumed that true survival rates would be included between the values estimated for such scenarios. Results We independently assessed the effect of predator control on five different parameters: (1) predator abundance; (2) survival of adult partridges; (3) nests success; (4) survival of Chicks_1; and (5) survival of Chicks_2.

Animal Biodiversity and Conservation 35.2 (2012)

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Table 3. Ranking of models resulting from the survival analysis with Program MARK for adult partridges. The model including the treatment (predator control) is underlined.: MLh. Model Lekelihood; Np. Number parameter; D. Deviance. Tabla 3. Ordenación de los modelos resultantes del análisis de supervivencia con el Programa MARK para perdices adultas. El modelo que incluye el tratamiento (control de depredadores) aparece subrayado: MLh. Modelo de probabilidad; Np. Número de parámetro; D. Desviación. Model

AICc

ΔAICc AICc weights

S4 Periods

259.515

0.213

251.448

S4 Periods+Age

260.111

0.596

0.158

0.742

250.009

S4 Periods+Year

260.429

0.914

0.134

0.633

250.327

S4 Periods+Study Area

261.528

2.013

0.077

0.365

252.426

S4 Periods+Treatment

261.545

2.030

0.077

0.362

251.443

S4 Periods+Sex

261.546

2.031

0.077

0.362

251.444

Constant Survival

262.655

3.139

0.044

0.208

260.648

Effects of predator control on predator abundance During the study, 143 foxes were harvested in the control zones (70 in 2008; 73 in 2009; table 2). During 2009, fox control had a clear effect on fox abundance, mainly in Study Area 1, where the KAI clearly decreased later in the year. In autumn, the fox KAI value was around 0.22 in the zones where control was applied, while it was 0.46 in zones without control (fig. 1A). In Study Area 2, although the effects were not so evident, the fox KAI in the NPC zone was almost always above the value of the KAI index in the PC zone (fig. 1B). In 2008, the low number of spotlight counts prevented a similar analysis. Even so, the only spotlight carried out, in August 2008, showed that fox abundance was lower in PC zones (fox KAI 0.19) than in NPC zones (fox KAI 0.26). Effect of predator control on survival of adult partridges Eighty–nine adult partridges were captured and radio– tagged during this study (52 in PC zones; 37 in NPC zones). Overall, 44% of the partridges were alive at the end of the tracking period in the PC zones, compared to 54% in the NPC zones. Predation was the main identified cause of death (68.2% of total deaths of radio–tagged animals, which represent 33.7% of the total number of radio–tagged partridges). Most adult predation (73%) occurred between April and June (30% in April; 30% in May and 13% in June). Predations caused by raptors occurred in April (54%) and May (46%), while predation by carnivores occurred mostly in May (58.3%) and June (33.3%), during the nesting period. However, predation by carnivores continued to occur until the end of the tracking period, although in lower proportions. The best model of adult survival included the time of year, and it was followed by the models that also

MLh

included age, year and study area (table 3). Although the model including the treatment (control / non–control) is not among the most plausible models, it can not be completely ruled out (ΔAIC = 2.030; table 2). The treatment had a much lower relative weight (0.077) to explain survival of adult partridges than the time of year (0.736), age (0.158) or year (0.134). Effect of predator control on nest survival We located 45 nests during the study (27 in PC zones; 18 in NPC zones). Thirty–three percent of the nests across all treatments had hatched at the end of the breeding season. Predation was identified as the main cause of nest loss in both study areas, accounting for 84% of all losses. Medium–size carnivores (mainly foxes, but also dogs and badgers) were identified as the main nest predators (30% of total nest predations; 39% in PC zones and 18 % in NPC zones). Mustelids and small mammals (hedgehogs, rats) were secondary nests predators (26.7% of total nest predations; 22.2% in PC zones and 33.3% in NPC zones). Corvids predated 3.3% of the nests. Other identified causes of nest loss were predation of the hen by raptors (6.6%), agriculture and livestock (6.6%) and nest abandonment by the adult (10%). It was not possible to identify the cause of 20% of nest losses, since no egg remains or evidence was found around the nest. In these cases predation was assumed to be the cause of nest loss. Four of the five models that best explained nest survival included the study area (table 4). The predator control treatment was included in the fifth ranked model, which also included the study area, with a ΔAIC of 1.927 (table 3). The model including only the study area explains nest survival better than the model which also included the treatment. However, no model was better (ΔAIC < 2) than the constant survival model.

400

Mateo–Moriones et al.

Table 4. Ranking of models resulting from the survival analysis with Program MARK for partridge nests. The model including the treatment (predator control) is underlined. (For abbreviations see table 3.) Table 4. Ordenación de los modelos resultantes del análisis de supervivencia con el Programa MARK para los nidos de perdiz. El modelo que incluye el tratamiento (control de depredadores) aparece subrayado. (Para las abreviaturas ver tabla 3.) Model

AICc

ΔAIC

AICc weights

MLh

Study Area

195.352

0.263

191.331

Study Area+Sex

195.791

0.439

0.211

0.802

189.750

Study Area+Year

196.600

1.248

0.141

0.535

190.559

Constant Survival

197.158

1.806

0.106

0.405

195.151

Study Area+Treatment

197.279

1.927

0.101

0.381

191.238

Study Area+Period

197.365

2.013

0.096

0.365

191.324

Sex

197.715

2.363

0.081

0.306

193.695

Among the factors considered, the relative weight of treatment (predator control) to explain nest success was much lower (0.101) than the study area (0.812), or the sex of the incubating parent (0.292). Effect of predator control on survival of Chicks_1 Sixty–two chicks between one and four days old (34 in PC zones; 28 in NPC zones) were captured and radio–tagged during the study. Overall, four chicks were radio–tracked until the transmitter battery went flat about 15 days after tagging (1 in PC zones and 3 in NPC zones), Eleven nests were undoubtedly predated (5 in PC zones and 6 in NPC zones). It was difficult to identify the nature of predation due to the small size of the chicks at this age: often, no remains or evidence other than the transmitter was werefound. The location of the transmitters and signs on it and on the surrounding vegetation revealed that at least three (27.3% of total predations in NPC zones) were predated by mustelids, two (18.2% of total in NPC zones) by foxes, and one (9.1%, in PC zones) by avian predators. We were not able to identify the predator in five cases (45.5%). Twenty–eight transmitters found showed no evidence of predation (21 in PC zones and 7 in NPC zones). Similar values of fallen tags were observed in previous tests in captivity (Mateo–Moriones et al., 2012). We lost the signal of 19 transmitters and, consequently, the final fate of those chicks was unknown. These chicks were considered in the two extreme scenarios previously mentioned (see Material and Methods). The best models for small chick survival under the minimum survival scenario were the constant model, and those including the treatment (ΔAIC = 0.373), the year, the study area and the age of the chicks (table 5A). The constant model had a relative weight of 0.320, while the model with the treatment had a relative weight of 0.262 under the minimum survival scenario. The best models under the maximum survival scenario were

those including the year and the constant model (table 5B). The model that included treatment was ranked in third place, with a ΔAIC = 2.043 (table 5B). Predator control treatment was the factor that best explained the survival of small chicks (relative weight: 0.168) in those factors included in the minimum survival scenario (year weight: 0.099) and the second best factor (0.105, after year, 0.292) under the maximum survival scenario. Estimated daily survival for the chicks during their first two weeks of life was slightly higher in the zones with predator control (between 0.961 ± 0.017 under the minimum and 0.974 ± 0.015 under the maximum survival scenario) than in the zones without predator control (between 0.918 ± 0.029 and 0.988 ± 0.011, respectively). Effect of predator control on survival of Chicks_2 We captured and radio–tagged 46 one–month– old chicks during the study, 35 in PC zones and 11 in NPC zones. A total of 24 chicks survived until the end of the radio–tracking period (November), 19 (54%) in the PC zones and five (45%) in the NPC zones. At least three chicks (9%) were predated in the PC zones, and two (18%) in the NPC zones. The remaining tagged birds (37% in both zones) lost their transmitters or their signal was lost during the tracking period. Under the minimum survival scenario, the model including the treatment as factor ranked best, above the constant model (Δ AIC = 1.461; table 6A), whereas under the maximum survival scenario, the model including the treatment ranked in second place after the constant model (ΔAIC = 0.491; table 6B). The predator control treatment was the factor that best explained the survival of large chicks both under the minimum survival (relative weight: 0.421) and under the maximum survival scenarios (relative weight: 0.178). Weekly survival rates for chicks over one–month old was estimated to be between 0.951 ± 0.017 and 0.987 ± 0.009 (for minimum and maximum survival

Animal Biodiversity and Conservation 35.2 (2012)

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Table 5. Ranking of models resulting from the survival analysis with Program MARK for partridge chicks in the first two weeks of life: A. Under the minimum survival scenario; B. Under the maximum survival scenario. The model including the treatment (predator control) is underlined. (For abbreviations see table 3.) Table 5. Ordenación de los modelos resultantes del análisis de supervivencia con el Programa MARK para pollos de perdiz durante las dos primeras semanas de vida: A. Bajo el escenario de mínima supervivencia; B. Bajo el escenario de máxima supervivencia. El modelo que incluye el tratamiento (control de depredadores) aparece subrayado. (Para las otras abreviaturas ver la tabla 3.) A Model

AICc

ΔAICc

AICc weights

MLh

Constant survival

89.964

0.203

87.945

Treatment

90.338

0.373

0.168

0.829

86.279

Year

91.396

1.432

0.099

0.488

87.338

Study Area

91.661

1.697

0.087

0.428

87.603

Year

91.728

1.764

0.084

0.414

87.670

Time

99.277

9.313

0.002

0.009

73.702

B Model

AICc

ΔAICc

AICc weights

MLh

Year

38.695

0.292

34.635

Constant survival

39.237

0.543

0.223

0.762

37.217

Treatment

40.737

2.043

0.105

0.360

36.678

Study Area

41.206

2.511

0.083

0.285

37.146

Age

41.253

2.558

0.002

0.278

37.193

Brood

47.892

9.19

0.000

0.010

26.752

scenarios, respectively) in the PC zones and between 0.867 ± 0.050 and 0.956 ± 0.031 in the NPC zones. Discussion In the present work predation was identified as the main cause of mortality for adults, nests and chicks of the red–legged partridge. Therefore, it would be expected that measures aimed to reduce predator abundances should increase partridge survival. Nevertheless, according to our results, predator control had different effects on each age group . It did not improve the survival of adults and nests, while it had a positive effect on chick survival, more evident in larger chicks. The lack of a decrease in fox abundance on the predation control sites early in the year, i.e. when nesting takes place, may explain the lack of an ability to show higher survival both of adults and of nests on the predation control plots. Differences in fox abundances appeared later in the year, coinciding with the end of the small chicks period and with the larger chicks period. Previous studies have reported improvements as high as 40% in nest success of red–legged partridges

through intensive control of foxes, magpies, dogs and feral cats (Yanes et al., 1998; Herranz, 2000). However, our study did not show such an improvement in the nest success due to predator control. The lack of differences in fox abundances between zones during the nesting period could explain this lack of effect. This suggests that to obtain results in the breeding season, fox control should start before that season. However, a high number of nest predations was due to other non–controlled predators, particularly some small mammalian predators, such as mustelids, hedgehogs and rodents. Unfortunately, we lack information on the abundance of such predators during the study. Corvids are usually considered important nest predators but they had a very low effect on our nests. This is probably associated with their low abundance in our study areas. In our study, predator control clearly improved survival rates for large chicks, under both scenarios considered. However, the survival of small chicks was only slightly improved by predator control despite predation being the most important cause of chick mortality. In a similar study conducted in south–central Norway, Steen & Haugvold (2009)

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Mateo–Moriones et al.

Table 6. Ranking of models resulting from the survival analysis with Program MARK for partridge chicks after the first month of life: A. Under the minimum survival scenario; B. Under the maximum survival scenario. The model including the treatment (predator control) is underlined. (For abbreviations see table 3.) Table 6. Ordenación de los modelos resultantes del análisis de supervivencia con el Programa MARK para pollos de perdiz tras el primer mes de vida: A. Bajo el escenario de mínima supervivencia; B. Bajo el escenario de máxima supervivencia. El modelo que incluye el tratamiento (control de depredadores) aparece subrayado. (Para las abreviaturas ver tabla 3.) A Model

AICc

ΔAICc

Treatment

103.428

Constant survival

104.889

Weight Year

AICc weights

MLh

0.421

99.371

1.461

0.203

0.482

102.871

104.952

1.524

0.196

0.467

100.894

106.271

2.842

0.102

0.241

102.213

Study Area

106.818

3.389

0.077

0.183

102.760

Time

116.865

13.437

91.281

B Model

AICc

ΔAICc

AICc weights

MLh

Constant survival

41.435

0.228

39.415

Treatment

41.926

0.491

0.178

0.782

37.867

Study srea

41.993

0.557

0.172

0.756

37.934

Weight

42.246

0.811

0.152

0.666

38.187

Year

43.254

1.819

0.092

0.403

39.195

Time

57.217

15.782

31.592

reported that local, intensive predator control had no measurable effects on chick production or survival of willow ptarmigan (Lagopus lagopus) even when predation was identified as the main cause of death. They suggested that their control areas may have been relatively small, which could have allowed the immigration of predators from local areas. It is possible that a similar effect may have taken place in our study, particularly in Study Area 2, where the decreasing trend of fox abundance was not quite clear throughout the experiment. A non–well conducted predator control carried on by the gamekeeper in this hunting estate, maybe less intensive, or extending also into the a priori non–predator control area, could explain the differences in the decreasing trend of fox abundances between the two study areas, but we have no data to confirm that –Amundson & Arnold (2011) observed that fox removal had no positive effect on mallard (Anas platyrhynchos) duckling survival, but this could be related to the high abundance of mink, which was not controlled. Similarly, we observed high predation by small carnivores, but their control was not considered in our study as they are protected in Spain. In a classical study in Southern England, predator control carried out on two hunting estates improved

brood size and abundance of the grey partridge (Perdix perdix, Tapper et al., 1996). Several differences between our work and that of Tapper et al. (1996) might explain the different results obtained in our study. Tapper et al. (1996) carried out intensive predator control over three consecutive years in each area in the study compared to just one year in our study. The effect of predator control was probably accumulative over the years, an effect that was not possible in our study. Furthermore, more species of predator were controlled in the English study, such as corvid species rather than magpies and some small mustelids. These predators were not controlled in our study (most are protected species in Spain), but they e had important roles as predators in our study area, mainly for nests, as they were was the second cause of losses, and probably for small chicks, even when the effect in this age group e was not easy to test (Calderón, 1977). Finally, there are some ecological differences between the two study areas, mainly related to biodiversity and predator abundance. A high diversity of predators, including mammals, birds and reptiles occur in the Iberian peninsula, and over 30 of these include red–legged partridge in their diets (Calderón, 1977; Duarte et al., 2008). Control was applied only to two of the predator species in our study and this

Animal Biodiversity and Conservation 35.2 (2012)

may not have been enough to reduce the effects of predation; it should be considered that the wide diversity of predator species in our study area includes several raptor species identified as partridge predators. Raptor predators are an important source of adult mortality, mainly during the mating period (Calderón, 1977; Buenestado et al., 2009). In conclusion, predator control, carried out in our study as performed in most Spanish hunting estates, was not effective in improving survival of adult partridges and their nests, probably because it was not effective in reducing abundances over a short period of time. Future research using indirect measures based on habitat improvements (nesting habitat, food and refuge availability) during the nesting season may prove effective to reduce partridge mortalities. A nest habitat with adjacent vegetation cover during the first days of life, for example, may reduce the need for the chicks to walk long distances looking for food (mainly small arthropods), and thus decrease the risk of being predated. In addition to well–designed selective predator control campaigns such as starting controls earlier, such measures could be effective to mitigate predation and improve survival of partridge populations. Acknowledgements This study was funded by the Department of Environment of Navarre Government and the Spanish Research Council (CSIC). We specially thank the staff of Viveros y Repoblaciones de Navarra for their support during the field work, and the Hunting societies of Sesma and Arróniz for their collaboration performing the predator management. The manuscript was greatly improved by the suggestions of the associate editor and an anonymous referee. References Aebischer, N. J. & Potts, G. R., 1994. Red–legged Partridge. In: Birds in Europe. Their conservation status: 214–215 (G. M. Tucker & M. F. Heath, Eds.). Birdlife International, Cambridge, UK. Amudson, C. L. & Arnold, T. W., 2011. The role of predator control removal, density–dependence, and environmental factors on mallard duckling survival in North Dakota. The Journal of Wildlife Management, 75: 1330–1339. Barilani, M., Bernard–Laurent, A., Mucci, N., Tabarroni, C., Kark, S., Garrido, J. A. P. & Randi, E., 2007. Hybridisation with introduced chukars (Alectoris chukar) threatens the gene pool integrity of native rock (A. graeca) and red–legged (A. rufa) partridge populations. Biological Conservation, 137: 57–69. Beja, P., Gordinho, L., Reino, L., Loureiro, F., Santos– Reis, M. & Borralho, R., 2009. Predator abundance in relation to small game management in southern Portugal: conservation implications. European Journal of Wildlife Research, 55: 227–238. Blanco–Aguiar, J. A., García, J. F., Ferreras, P., Viñuela, J. & Villafuerte, R., 2001. Effect of game

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management on artificial nest predation in central Spain. In 25th International Union of Game Biologists (IUGB) and the 9th International Symposium Perdix, Limasol, Chipre. Blanco–Aguiar, J. A., González–Jara, P., Ferrero, M. E., Sánchez–Barbudo, I., Virgós, E., Villafuerte, R. & Dávila, J. A., 2008. Assessment of game restocking contributions to anthropogenic hybridization; the case of the Iberian Red–legged Partridge. Animal Conservation, 11: 535–545. Blanco–Aguiar, J. A., Virgos, E. & Villafuerte, R., 2003. Perdiz Roja (Alectoris rufa). In: Atlas de las aves reproductoras de España: 212–213 (R. Marti & J. C. del Moral, Eds.). Dirección General de Conservación de la Naturaleza y Sociedad Española de Ornitología, Madrid, Spain. Buenestado, F. J., Ferreras, P., Delibes–Mateos, M., Tortosa, F. S., Blanco–Aguiar, J. A. & Villafuerte, R., 2008. Habitat selection and home range size of red–legged partridges in Spain. Agriculture Ecosystems & Environment, 126: 158–162. Buenestado, F. J., Ferreras, P., Blanco–Aguiar, J. A., Sánchez–Tortosa, F. & Villafuerte, R., 2009. Survival and causes of mortality among wild Red– legged Partridges Alectoris rufa in southern Spain: implications for conservation. Ibis, 154: 720–730. Calderón, J., 1977. El papel de la Perdiz roja (Alectoris rufa) en la dieta de los predadores ibéricos. Doñana, Acta Vertebrata, 4: 61–126. Casas, F., Mougeot, F. & Viñuela, J., 2009. Double– nesting behaviour and sexual differences in breeding success in wild Red–legged Partridges Alectoris rufa. Ibis, 151: 743–751. Casas, F. & Viñuela, J., 2010. Agricultural practices or game management: which is the key to improve red–legged partridge nesting success in agricultural landscapes? Environmental Conservation, 37: 177–186. Cooch, E. & White, G., 2010. Program Mark: A gentle introduction. Online Access: http://www.phidot.org/ software/mark/docs/book/. Côté, I. M. & Sutherland, W. J., 1997. The effectiveness of removing predators to protect bird populations. Conservation Biology, 11: 395–405. Cramp, S. & Simmons, K. E. L., 1980. Handbook of the birds of Europe, the Middle East and North Africa. Oxford Univ. Press, Oxford, London & New York. Crooks, K. R., & Soulé, M. E., 1999. Mesopredator release and avifaunal extinctions in a fragment system. Nature, 400: 563–566. Delibes–Mateos, M., Ferreras, P. & Villafuerte, R., 2009. European rabbit population trends and associated factors: a review of the situation in the Iberian Peninsula. Mammal review, 39: 124–140. Díaz–Sanchez, S., Mateo–Moriones, A., Casas, F. & Höfle, U., 2011. Prevalence of Escherichia coli, Salmonella sp. and Campylobacter sp. in the intestinal flora of farm–reared, restocked and wild red–legged partridges (Alectoris rufa): is restocking using farm–reared birds a risk? European Journal of Wildlife Research, DOI 10.1007/ s10344–011–0547–5. Duarte, J., Farfan, M. A. & Guerrero, J. C., 2008.

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Importancia de la predación en el ciclo anual de la perdiz roja. In: Especialista en control de predadores: 133–141 (J. L. Garrido, Ed.). FEDENCA, Alcobendas, Madrid. Herranz, J., 2000. Efectos de la depredación y del control de predadores sobre la caza menor en Castilla–La Mancha. Ph. D. Thesis, Univ. Autónoma de Madrid. Mateo–Moriones, A., Villafuerte, R. & Ferreras, P., 2012. Evaluation of radiotagging techniques and their application to survival analysis of Red-legged Partridge Alectoris rufa chicks. Ibis, 154: 508–519. Millán, J., Gortázar, C. & Villafuerte, R., 2001. Marked differences in the splanchnometry of farm–raised and wild red–legged partridges (Alectoris rufa). Poultry Science, 80: 972–975. Moleón, M., Almaraz, P. & Sánchez–Zapata, J. A., 2008. An emerging infectious disease triggering large–scale hyperpredation. PLoS ONE, 3: e2307. Moynahan, B. J., Lindberg, M. S., Rotella, J. J. & Thomas, J. W., 2007. Factors affecting nest survival of greater sage–grouse in North Central Montana. Journal of Wildlife Management, 71: 1773–1783. Negro, J. J., Torres. M. J. & Godoy, J. A., 2001. RAPD analysis for detection and eradication of hybrid partridges (Alectoris rufa x A. graeca) in Spain. Biological Conservation, 98: 19–24. Rotella, J. J., Dinsmore, S. J. & Shaffer, T. L., 2004. Modeling nest–survival data: a comparison of recently developed methods that can be imple-

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Effectiveness of habitat management for improving grey partridge populations: a BACI experimental assessment E. Bro, P. Mayot & F. Reitz

Bro, E., Mayot, P. & Reitz, F., 2012. Effectiveness of habitat management for improving grey partridge populations: a BACI experimental assessment. Animal Biodiversity and Conservation, 35.2: 405–413. Abstract Effectiveness of habitat management for improving grey partridge populations: a BACI experimental assessment.— We assessed the impact of field division (4 m bare ground strips within wheat fields) and food supplementation (supplied through grain feeders) on grey partridge Perdix perdix L. populations using six–year 'before–after'/'control–impact' (BACI) experiments. We did not detect any convincing positive effects of either of these two schemes on partridge pair density and reproductive success. Increases in pair densities were similar on managed and control areas, and contrasting results were found between some sites. No consistent pattern was observed between reproductive success and feeding intensity. Our studies highlight the need for field experiments at farm–scale to test the effectiveness of management measures. We conclude that, in the context in which they are applied, management techniques directed towards increasing partridge density do not systematically provide the desired outcome. We develop our point of view about management in the Discussion. Key words: BACI experiments, Farm–scale, Grey partridge, Habitat management, Reproductive success, Spring density. Resumen Eficacia de la gestión del hábitat para mejorar las poblaciones de perdiz pardilla: una evaluación experimental BACI.— Evaluamos el impacto de la división de los campos (franjas de 4 m de suelo desnudo dentro de campos de trigo) y de la alimentación suplementaria (mediante suministradores de grano) sobre la perdiz pardilla Perdix perdix L., utilizando experimentos antes–después/control–impacto (BACI, 'before–after'/'control–impact') de seis años. No detectamos ningún efecto positivo convincente de ninguna de estas dos medidas sobre la densidad de parejas de perdices ni el éxito reproductivo. Los aumentos en la densidad de parejas fueron similares en las áreas de control y en las gestionadas y se encontraron resultados contrastantes entre algunos emplazamientos. No se observó ningún patrón consistente entre el éxito reproductivo y la intensidad de la alimentación. Nuestros estudios destacan la necesidad de experimentos de campo en granjas, para comprobar la eficacia de las medidas de gestión. Nuestra conclusión es que, en el contexto en que se aplicaron, las técnicas de gestión dirigidas a aumentar la densidad de perdices no produjeron sistemáticamente el efecto deseado. En la Discusión desarrollamos nuestro punto de vista sobre la gestión. Palabras clave: Experimentos BACI, Granja, Perdiz pardilla, Gestión del hábitat, Éxito reproductivo, Densidad primaveral. Received: 27 I 12; Conditional acceptance: 11 V 12; Final acceptance: 26 VI 12 Elisabeth Bro, Pierre Mayot & François Reitz, National Game & Wildlife Agency (ONCFS), Research Dept., Sedentary Small Game Team, BP 20, 78 612 Le Perray–en–Yvelines cedex, France. Corresponding author: Elisabeth Bro. E–mail: elisabeth.bro@oncfs.gouv.fr

ISSN: 1578–665X

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Introduction The decline in grey partridge Perdix perdix L. populations after the Second World War has mainly been attributed to farming intensification and related loss of habitat quality (loss of crop diversity, field enlargement, scarcity of cover and food resources after harvest, etc., see the recent review of Kuijper et al., 2009). This is assumed to have led to a limitation in resources such as food and nesting sites. As a consequence, habitat management is often recommended to improve the carrying capacity of grey partridge on hunting estates. However, only a few studies have tested the effectiveness of the tools available for this on a wide scale. From 2000–2007, we ran three projects examining the impact of various habitat management tools on grey partridge populations with the aim of having experimental verification, as well as, in case of positive results, demonstration sites for local hunting associations to motivate hunters and farmers to apply some management techniques more widely and intensively. We separately assessed the effects of: (1) wildlife cover after cereal harvest, using maize–sorghum strips; (2) food supplementation, provided through grain feeders; and (3) dividing cereal fields, seeking to increase nesting cover within the cereal ecosystems, by using 4m bare ground strips to divide large cereal fields, using 'before–after'/'control–impact' (BACI) experiments. We chose to examine these simple technical measures separately to dissociate their effect from the global effect of the usual package of management recommendations which combine habitat management, food supplementation and predator control. The two first measures are currently commonly applied and the third one could reasonably be applied. In this paper, we report the results of the last two experiments; the first one has already been published in detail (see Bro et al., 2004). Hereafter we present the context of the studies. Exp. 2. Food supplementation Grey partridge densities vary from low (< 5 pairs/100 ha) to high (30–40 pairs/100 ha, or even more locally) levels in central–northern France (e.g. Bro & Crosnier, 2012; Bro et al., 2005; Mérieau & Bro, 2009). In the late 1990s, we carried out an inquiry to identify and quantify the management techniques that were applied on managed hunting estates in this region (Mayot, 1999; 485 estates totaling 502,000 ha). Supplementary feeding appeared to be the most widely applied measure, occurring on 93.6% of estates. Other measures included plantation of hedges (on 41% of managed hunting estates, abundance varying between 3 and 1,600 m/100 ha); game cover (62%, < 0.10–ca. 10% of arable land), predator control (78.5%; judged as light on 22.5% estates, moderate on 28.4% and intensive on 27.8%). Furthremore, supplementary feeding was the only measure applied on a quarter of estates. The reason for this is that even though feeding is costly and time consuming, it can be easily applied by hunters who are not involved in farming the estate but are trying to improve habitat. Their ultimate aim

Bro et al.

is to increase partridge stock in spring and improve reproductive success. The mechanisms involved are believed to be both a reduction in dispersal rates in late winter and the improvement of nesting hen condition. However, feeding appeared to be extensive with 40.4% of managed hunting estates with ≤ 5 feeders/100 ha and 76.4% with ≤ 10 feeders/100 ha. Such application of the measure raises the question of its effectiveness given it often does not match the rule of 'one feeder for one pair' (fig. 1A) that is commonly recommended, all the more that we did not detect a positive relationship. No relationship was detected with the reproductive success (fig. 1B). The objective of our study was to test the impact of a more intensive feeding regime (feeder density was ≥ 20 feeders/100 ha on 5.3% of the 485 estates) than that routinely applied on managed hunting estates across France. Exp. 3. Field size division The context of this experiment was quite different from the previous study. Hunters of an estate were applying the recommended partridge conservation measures, combining feeding, habitat management, predator control and limitation of hunting bag. Despite the fact that partridge density was higher on their estate than on surrounding estates, they considered it could be even higher. A limited nesting carrying capacity was a possible explanation. As these hunters were also farmers, they accepted to divide their fields of cereals as a further habitat management strategy. Cereal edge near a lane is the preferred nesting habitat of the grey partridge in cereal ecosystems in France (Bro et al., 2000a; Reitz et al., 2002) and the habitat where most nests survive (Bro et al., 2000b). The willingness of these farmers to increase partridge density on their farms offered us the opportunity to test the effects of a measure that had only been examined indirectly previously, by comparing contrasting natural situations (see Bro et al., 2008). Methods Study sites The studies were carried out in the Beauce region, near the cities of Chartres (field size reduction experiment) and Orléans (supplementary feeding experiment). The Beauce region was the species’ core area in the 1980s, but partridge densities here experienced a marked decline during the 1990s and 2000s (see Bro et al., 2005; Mangin, 2009). Crop cover consisted of approximately 60% of cereals, 15% of sugar beet and ca. 5–10% of rapeseed and maize. Peas, sunflower and potatoes were the other cultivated crops. The landscape was typical of the region, with open fields separated by some groves but almost no hedges. Satellite maps are available at e.g. http://www.maplandia.com. Population survey Spring censuses were carried out to estimate the grey partridge breeding stock. Counts were performed in

Animal Biodiversity and Conservation 35.2 (2012)

A Partridge density

80 70 60 50 40 30 20 10

Reproduction success

0 B

407

10 8 6 4 2 0

20 30 40 50 Density of feeders (feeders/100 ha)

Fig. 1. Relationship between feeder density and: A. Partridge density (pairs/100 ha); B. Reproductive success (offspring/female) across managed hunting estates. (Unpublished data from wild populations in central–northern France.) Fig. 1. Relación entre la densidad de alimentadores y: A. Densidad de perdices (parejas/100 ha); B. Éxito reproductivo (crías/hembra). (Datos no publicados de las poblaciones salvajes en el centro septentrional de Francia .)

March, when birds had paired and before crop cover was too high. The census was obtained by counting the number of partridges flushed while fields were beaten by a line of people (units called 'beats'; fig. 2). A full description of the field procedures is given in Bro et al. (2005). The same method was used on all sites, all areas (experimental and control), and all years.

On Bougy–Neuville, it was increased by 2.6 (from 7.7 to 20; n = 193). Thus we tested the effects of an increase in feeding intensity, compared to the baseline application of the measure. Feeders were static during the course of the study and their number did not change. The total use of wheat grain was roughly estimated to 11–12 T/ year. The experiment cost ca. 10 k€.

Exp. 2. Supplementary feeding experiment

Exp. 3. Field size reduction experiment

The experiment was replicated on two sites of 420 ha ('Oison') and 990 ha ('Bougy–Neuville'). Cereals crops amounted to 70% and 65% of arable land, respectively; rapeseed 15% and 14%, maize < 1% and 6.4%, permanent meadows 0% and 3.8%. Intensive feeding started in autumn 2003 and finished in autumn 2006. Feeding occurred from September to June. Feeders were mostly located along lanes; sometimes between wheat and sugar beet or maize fields. The density of feeders ranged locally between 10 and 50/100 ha across the experimental area, depending upon partridge density on beats in spring 2003 (with the rule of ca. 1 feeder/bird). On Oison, we increased mean feeder density 7 fold (from 6.1 to 43.6 feeders/100 ha, n = 177 during the experiment).

It was only possible to carry out this experiment on one site of 600 ha ('Aubepine'). As explained in the Introduction, the site was not chosen randomly. The area of winter wheat amounted to 53% of arable land and winter barley 9%. Mean field size of winter wheat was ca. 10 ha (12 ha in 2003, 9 ha in 2004 and 10 ha in 2005), varying from 3 to 41 ha (see a map in Mayot et al., 2009b). The experimental site was managed for partridges and pheasants. Wildlife set–aside was planted on 5% of arable land and bushes on 1%. In autumn, the hunting bag varied from 7 to 15 partridges/100 ha depending upon partridge density and reproductive success; it was achieved through 1 or 2 hunts. Legal predator control limited the number of red foxes, stone martens, carrion crows and magpies on

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Chartres

C 500 m

Orléans Experimental area

20 km

Adjacent control area Non–adjacent control area

N S

Fig. 2. Location of study sites within: A. France; B. Relative location of communalities with experimental (black) vs. control (grey) areas; and C. Spatial design of beats (ca. 80–150 ha) where censuses are carried out. In the example of Oison (C), the dotted line represents the limit of the experimental area, monitored through three beats. Adjacent beats are in dark grey and non–adjacent beats in light grey. Fig. 2. Localización de los lugares de estudio: A. Francia; B. Localización relativa de las comunalidades con áreas experimentales (en negro) vs. áreas de control (en gris); C. Diseño espacial de las batidas (aprox. 80–150 ha) donde se llevaron a cabo censos. En el ejemplo de Oison (C): la línea de puntos representa el límite del área experimental, que fue monitorizada mediante tres batidas. Las batidas adyacentes están en gris oscuro, y las no adyacentes en gris claro.

the site (bag not available). Wheat grain was provided through ca. 100 feeders. We collaborated with local farmers to reduce the overall size of winter wheat fields by using 4 m strips of bare ground to divide the fields. We tested strips of bare ground rather than strips of game cover because it is a simple technique which did not require any additional farming operation since they contained no cover. As the strips corresponded to an area where the grain was not sown they were managed in the same way as the crop except for insecticide and fungicide spraying. Strips were not located in winter barley fields because, in this region, this crop is harvested in late June, coinciding with chick hatching. Approximately 20 strips were introduced (23 in 2003, 24 in both 2004 and 2005), corresponding to a total area of 3 ha (3.291, 3.266 and 2.973 ha in 2003, 2004 and 2005, respectively) and a total length of 8km (8.80, 8.87 and 7.82 km). The mean size of winter wheat fields during the experiment was reduced by 1.5 (data of 2004). The location of strips varied from year to year depending upon crop rotation. These strips represented

an additional abundance of wheat margins of 60 m/ha of winter wheat (61.5, 77.8, 58.5 m/ha). The increase amounted to 20% of the initial level. A compensation of 762 € / ha was paid both for yield loss and the lack of CAP subsidies. The total cost amounted to 7.3 k€ for the 3 years. Experimental design To test the impact of a given management scheme as rigorously as possible, we conducted 6–year 'before–after'/'control–impact' (BACI) experiments. We used the BACI design to attempt to overcome the problem of ascribing changes to the scheme rather than natural variability. We replicated the study over 2 sites where possible and used several control areas to provide further robustness to our results, allowing spatial heterogeneity to be taken into account. Control areas were neighbouring areas (< 10 km, see fig. 2) so that habitat characteristics, weather, and predator abundance were assumed to be reasonably similar. Experimental and control areas were included in

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Table 1. Intensive feeding experiment: spring density (pairs/100 ha) in 2000–2003 (period 'before') and 2004–2006 (period 'after') and difference (%) between the two periods. Tabla 1. Experimento de alimentación intesiva: densidad primaveral (parejas/100 ha) en 2000–2003 (periodo "anterior") y 2004–2006 (periodo "posterior"), y diferencia (%) entre los dos periodos.

Before Area (ha)

2001–2003

After 2004–2006

% difference before/after

Site 1: Oison Experimental area

420

16.5

24.5

+48.5

Adjacent control area

260

11.9

23.7

+100

Non–adjacent control area

500

–31.3

Experimental area

990

11.7

25.9

+121.4

Adjacent control area

320

9.8

17.7

+80.6

Non–adjacent control area

820

10.8

27.7

+156.5

20.4

+77.4

Site 2: Bougy–Neuville

Other control areas (neighbouring municipalities of the same GIC) Saint–Lyé

500

11.5

Santeau

490

14.6

31.5

+115.8

Aschères

640

15.8

21.2

+34.2

Chilleurs

620

14.6

16.8

+15.1

Crottes

580

+63.6

Montigny

370

16.1

26.8

+66.5

the same 'GIC' (i.e. grouping of several hunting estates to share a same game management scheme over an area of several thousands of ha), so that the baseline management could be considered as reasonnably similar as well. It has not changed during the course of the studies, except for the manipulated factor on the experimental area. We distinguished two kinds of control areas in the feeding experiment depending upon whether beats were adjacent (boundaries < ca. 300 m) or not to the experimental area (see fig. 2).The tables also provide data of partridge density on all surrounding municipalities from the same GIC where partridge populations have been routinely surveyed on a long– term basis. Field data of these additive control areas were extracted from the database of the national grey partridge population survey (see Bro et al., 2005). Data analysis Experiments were carried out on a a large scale, and although we tried to replicate them (as we did for the ‘cover’ experiment), available data did not allow us to reasonably use the same statistical tests as in Bro et al. (2004). Instead we tested the relationship between the mean density or the mean reproductive success and the period ('before' vs. 'after') * area ('experimental' vs. 'control') interaction using an ANOVA (proc GLM, type III, SAS Institute). Year and area were included

as co–variables. Reproductive success was tested against the intensity of feeding (feeder–to–pair ratio) during the 'after' period using a glm with year and beat as co–variables (proc GLM). Results Impact of intensive feeding on spring density On Oison, we observed an overall increase in densities on the 'feeding' area (table 1, fig. 3A). The rate of increase was higher than that observed on three of the eight control areas. Statistically, this increase was not related to intensive feeding (P = 0.520). The pattern was quite different on Bougy–Neuville where feeding was twice as intensive (table 1, fig. 3B). Mean partridge density was stable from 2001 to 2004 and increased notably in the last two years, i.e. with a time lag of one year after started feeding (fig. 3B). Qualitatively, the same pattern was observed on both the adjacent and the non–adjacent areas (fig. 3B). Quantitatively, the increase rate was higher on the non–adjacent control area and lower on the adjacent area (table 1), but overall the difference was not statistically significant (P = 0.741). In addition, similar increases were observed in the other control areas (table 1). No significant correlation was found at the local scale between the reproductive success and feeding intensity (P = 0.109, fig. 4).

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A Density (pairs/100 ha)

50 40 30 20 10 0 2000

50 Density (pairs/100 ha)

Experimental (420 ha) Non–adjacent control (500 ha) Adjacent control (260 ha)

Before

2001

2002

After

2003

2004

2005

2006

2007

Experimental (990 ha) Non–adjacent control (820 ha) Adjacent control (320 ha)

30 20 10 0 2000

Before

2001

2002

After

2003

2004

2005

2006

2007

Fig. 3. Intensive feeding experiment: changes in spring pair density depending upon whether intensive feeding is undertaken or not, in the managed vs. control areas. Vertical bars indicate min. and max. values of densities recorded across beats; they are provided to describe spatial variability: A. Oison; B. Bougy–Neuville. Fig. 3. Experimento de alimentación intensiva: cambios en la densidad primaveral de parejas, dependiendo de si se ha llevado a cabo la alimentación intensiva o no, en las áreas de gestión vs. las áreas de control. Las barras verticales indican los valores mínimos y máximos de las densidades registradas mediante las batidas; se han incluido para describir la variabilidad espacial: A. Oison; B. Bougy–Neuville.

Impact of wheat field size reduction on spring density We observed an increase in pair density in 2003–2005 compared to 2000–2002 in both the experimental and the control areas (table 2, fig. 5A), and this was not be attributable to our experiment (P = 0.562). No differential effect was detected on reproductive success (P = 0.403, fig. 5B). Unfortunately, we were unable to replicate our experiment at other sites. Discussion Our three experiments testing the impact of wildlife cover (Bro et al., 2004), intensive feeding (see further technical details in Mayot et al., 2009a) and field size division (Mayot et al., 2009b) did not provide

convincing, definitively positive effects in the short term. Perhaps the limiting factors on our study sites were not food and nesting sites. Our experiments coincided with several years of good grey partridge reproduction throughout France (see Reitz & Mayot, 2009), which might have contributed to our inability to identify any positive effects. However, from the great body of research that has been dedicated to this species, it is well known that partridge populations are influenced by multiple external (i.e. environmental) and intrinsic (e.g. density–dependence) factors that are likely to vary in space and time and to depend upon population status. Hence, several mechanisms might explain our results but we have so far not been able to identify them. Therefore, we do not conclude that the measures we experimentally tested are in inefficient but that, applied in the context described above, they

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175 125 75 25 –25 –75

Reproductive success (offspring / female)

Change in density (%)

411

8 6 4 2 0

2 3 Feeder–to–pair ratio

Fig. 4. Intensive feeding experiment: relationships between the changes in spring pair density (A) and the reproductive success (B) and the feeder–to–pair ratio during the three years of the 'after' period. Oison, black–filled symbol; Bougy–Neuville, open symbols. Each symbol corresponds to a separate beat. Note that coveys were not mapped on the Bougy–Neuville site so that figure B could not be drawn. Fig. 4. Experimento de alimentación intensiva: relación entre los cambios de la densidad primaveral de parejas (A) y el éxito reproductivo (B) y la tasa de alimentador–pareja durante tres años del periodo “after“, después. Oison, símbolos en negro; Bougy–Neuville, símbolos en blanco. Cada símbolo corresponde a una batida distinta. Nótese que no se mapearon las nidadas en el emplazamiento de Bougy–Neuville, de forma que no pudo dibujarse la figura B.

did not improve populations. This is compatible with previous experiments carried out at an individual level using radiotracking , which also moderated the impact of feeding on survival and reproduction, showing positive, negative or no effects (e.g. Haines et al., 2004; Hoodless et al., 1999; Townsend et al., 1999;

Valkeajärvi & Ljäs, 1994). In other words, our feeling is that all efforts do not always guarantee results. An important message we convey to hunters is to make a preliminary diagnosis of the characteristics of their estates to identify their actual weaknesses and then to focus on measures to counteract these (Bro

Table 2. Field size reduction experiment. (For more details see table 1.) Tabla 2. Experimento de reducción del tamaño del campo. (Para más detalles ver tabla 1.)

Area (ha)

Before 2000–2002

After 2003–2005

% difference before/after

Experimental area

600

23.3

31.7

+36.1

Non–adjacent control area (same GIC)

1110

10.6

20.6

+94.3

Other control area (neighbouring GIC)

1800

9.5

+26.3

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Bro et al.

Experimental (600 ha) Non–adjacent control (1,110 ha)

Density (pairs/100 ha)

40 30 20 10 0 1999

Before

2000

2001

2002

Reproductive success (offspring/female)

After

2003

2004

2005

Experimental area Other control area

56 34

Before

1999

2000

53 19

12 15

2006

2001

2002

After

2003

2004

2005

2006

Fig. 5. Field size reduction experiment: A. Changes in spring pair density depending upon whether intensive feeding is undertaken or not, in the managed vs. control areas. Vertical bars indicate min. and max. values of densities recorded across beats; they are provided to describe spatial variability; B. Reproductive success during the course of the experiment on the experimental vs. other control area. Figures indicate covey numbers. Fig. 5. Experimento de reducción del tamaño del campo: A. Cambios en la densidad primaveral de parejas, dependiendo de si se ha llevado a cabo la alimentación intensiva o no, en las áreas de gestión vs. las áreas de control. Las barras verticales indican los valores mínimos y máximos de las densidades registradas mediante las batidas; se han incluido para describir la variabilidad espacial; B. Éxito reproductivo durante el transcurso del experimento en el área de experimentación vs. área de control. Las cifras indican el número de batidas.

et al., 2009). We encourage hunters to improve the characteristics of their estates step by step, learning from a trial and error approach. Several partridge restoration programs have provided very good results (e.g. Mérieau & Bro, 2009; Connor & Draycott, 2010). This should encourage other hunters. A fully integrated management programme —including predator control, feeding and habitat management— was applied in these cases. The ultimate question is to what extent this can be done over a wide area and over the long term.

Acknowledgements This work was conducted in collaboration with field technicians of two hunter associations ('Loiret' and 'Eure–et–Loir'). We are grateful to all the people who collected the data. We acknowledge the landowners and hunters for giving us permission to work on their land and for their cooperation in the study. We acknowledge two anonymous reviewers and the editor for their criticisms on a previous draft of the paper that allowed us to improve it.

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References Bro, E. & Crosnier, A., 2012. Grey Partridges Perdix perdix in France in 2008: distribution, abundance, and population change, Bird Study, 59: 320-326. Bro, E., Mayot, P., Corda, E. & Reitz, F., 2004. Impact of habitat management on grey partridge population dynamics on cereal ecosystems in France: assessing a wildlife cover scheme using a multi–site BACI experiment. Journal of Applied Ecology, 41: 846–857. Bro, E., Mayot, P., Millot, F. & Reitz, F., 2009. A propos de l’aménagement de l’habitat pour la perdrix grise de plaine. Réflexion entre théorie et pratique. Faune Sauvage, 283: 28–31. Bro, E., Meynier, F., Sautereau, L. & Reitz, F., 2008. Peut–on prédire les densités de perdrix grise dans les plaines de grande culture? Faune Sauvage, 282: 26–34. Bro, E., Reitz, F., Clobert, J. & Mayot, P., 2000a. Nest–site selection of grey partridge (Perdix perdix) on agricultural lands in north–central France. Game and Wildlife Science, 17: 1–16. – 2000b. Nesting success of grey partridges (Perdix perdix) on agricultural land in North–Central France, relation to nesting cover and predator abundance. Game and Wildlife Science, 17: 199–218. Bro, E., Reitz, F. & Landry, P., 2005. Grey partridge population status in central northern France: spatial variability in density and 1994–2004 trend. Wildlife Biology, 11: 287–298. Connor, H. E. & Draycott, R. A. H., 2010. Management strategies to conserve the grey partridge: the effect on other farmland birds. Aspects of Applied Biology, 100: 359–363. Haines, A. M., Hernandez, F., Henke, S. E. & Bingham, R. L., 2004. Effects of road baiting on home range and survival of southern bobwhites in southern Texas. Wildlife Society Bulletin, 32: 401–411. Hoodless, A. N., Draycott, R. A. H., Ludiman, M. N. & Robertson, P. A., 1999. Effects of supplemen-

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tary feeding on territoriality, breeding success and survival of pheasants. Journal of Applied Ecology, 36: 147–156. Kuijper, D. P. J., Oosterveld, E. & Wymenga, E., 2009. Decline and potential recovery of the European grey partridge (Perdix perdix) population – a review. European Journal of Wildlife Research, 55: 455–463. Mangin, E., 2009. Situation de la perdrix grise en Eure–et–Loir. L’ancien bastion souffre… Faune Sauvage, 286: 37–40. Mayot, P., 1999. Aménagements pour la perdrix : résultats d’une enquête régionale. Bulletin Mensuel de l’ONC, 249: 28–32. Mayot, P., Malécot, M., Vigouroux, L. & Bro, E., 2009a. L’agrainage intensif : quel impact sur la perdrix grise? Résultats d’expérimentation en plaine de grande culture. Faune Sauvage, 283: 32–39. Mayot, P., Sautereau, L., Reitz, F. & Bro, E., 2009b. Division du parcellaire agricole et nidification de la perdrix grise. Résultats d’expérimentation en Beauce. Faune Sauvage, 283: 40–43. Mérieau, A. & Bro, E., 2009. Gestion de la perdrix grise dans les Ardennes: 25 ans d’efforts, des densités records. Faune Sauvage, 283: 44–50. Reitz, F., Le Goff , E. & Fuzeau, M., 2002. Landscape selection by grey partridge (Perdix perdix) for nesting in the fields of french cereal agrosystems. Game and Wildlife Science, 19: 209–220. Reitz, F. & Mayot, P., 2009. Le Réseau 'Perdrix– Faisans'. La situation des perdrix et faisans en 2008 dans le Centre–Nord de la France. Faune Sauvage, 283: 63–65. Townsend, D. E., Lochmiller, R. L., De Maso, S. J., Leslie, D. M., Peoples, A. D., Cox, S. A. & Perry, E. S., 1999. Using supplémental food and its influence on survival of nothern bobwhite (Colinus virginiatus). Wildlife Society Bulletin, 27: 1074–1081. Valkeajärvi, P. & Ljäs, L., 1994. Comparison of breeding success between fed and unfed black grouse in Central Finland. Suomen Riista, 40: 98–109.

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The invertebrate diet of northern bobwhite chicks in Georgia, United States D. A. Butler, W. E. Palmer & M. P. Cook

Butler, D. A., Palmer, W. E. & Cook, M. P., 2012. The invertebrate diet of northern bobwhite chicks in Georgia, United States. Animal Biodiversity and Conservation, 35.2: 415–418. Abstract The invertebrate diet of northern bobwhite chicks in Georgia, United States.— The establishment of brood–rearing habitats along field margins has become a popular agri–environmental prescription to help reverse population declines of northern bobwhite (Colinus virginianus) in Georgia, United States. Here, the invertebrate–diet of chicks foraging on farmland with established brood–rearing habitats is examined and compared to those of chicks on an intensively managed wild bobwhite shooting estate. In 2001 and 2002, faecal samples were collected and analysed from nocturnal roost sites of bobwhite broods. Differences in invertebrate composition between the study sites were investigated using compositional analysis. While the diet of chicks on both sites contained similar invertebrate groups, the composition of the diets varied significantly. Although chicks on farmland had eaten 1.7 times fewer Coleoptera, they had 1.7 times more Hemiptera in their diet. These data suggest that although the invertebrate composition in the diet of chicks differed between the two landscapes, both contained high proportions of important prey items. Key words: Northern bobwhite, Chick–diet, Invertebrates, Agri–environmental scheme, Brood–rearing habitat. Resumen Dieta a base de invertebrados de las crías del colín de Virginia en Georgia, Estados Unidos.— El establecimiento de hábitats de cría a lo largo de los márgenes de los campos se ha convertido en una norma agro–medioambiental muy popular, para ayudar a invertir la disminución de las poblaciones del colín de Virginia (Colinus virginianus) en Georgia, Estados Unidos. En este estudio se examina la dieta a base de invertebrados de los pollos, con hábitats de cría bien establecidos, que forrajean en las tierras de labrantío, en comparación con la de las crías de un coto de caza de colines de Virginia salvajes gestionado intensivamente. En los años 2001 y 2002 se recogieron muestras fecales de los lugares de descanso nocturnos de las crías, y se analizaron. Se investigaron las diferencias en cuanto a la composición de invertebrados entre los lugares de estudio, utilizando un análisis composicional. Mientras que la dieta de los pollos de ambos lugares contenía grupos de invertebrados similares, la composición de las dietas variaba significativamente. A pesar de que las crías de los cultivos habían comido 1,7 veces menos coleópteros, habían devorado 1,7 veces más hemípteros. Estos datos sugieren que aunque la composición de invertebrados de la dieta de las crías difería entre los dos tipos de hábitat, en ambos contenía grandes proporciones de las presas más importantes. Palabras clave: Colín de Virginia, Dieta de las crías, Invertebrados, Proyecto agro–medioambiental, Hábitat de cría. Received: 23 II 12; Conditional acceptance: 8 V 12; Final acceptance: 6 VII 12 David A. Butler & William E. Palmer, Tall Timbers Research Station and Land Conservancy, 13093 Henry Beadel Drive, Tallahassee, Florida 32312, USA.– M. P. Cook, Warnell School of Forestry and Natural Resources, The Univ. of Georgia, 180 E Green Street, Athens, Georgia 30602, USA. Corresponding author: D. A. Butler, Perdix Wildlife Solutions Ltd., Avenue R, Stoneleigh Park, Kenilworth, Warwickshire CV8 2LG, U.K. E–mail: dbutler@perdixwildlife.co.uk

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Introduction During the past 50 years northern bobwhite (Colinus virginianus) (hereafter bobwhite) populations have declined rapidly in the southeastern United States (Brennan, 1991). Declines have been notable on farmland where agricultural intensification has led to a reduction in habitat for bobwhite (Brennan, 1991). Of particular concern is the loss of brood–rearing habitats that harbour high densities of invertebrates important in the diet of chicks (Stromborg, 1982). Loss of quality brood–rearing habitat can prevent game bird chicks from obtaining sufficient quantities of prey items and consequently reduce survival rates (Potts, 1986). The daily number of prey–items required by gamebird chicks depends upon the size and nutritional value of the invertebrates within the diet (Southwood & Cross, 2002). In grey partridge, Perdix perdix, for example, Southwood & Cross (2002) reported that a nine–day–old grey partridge chick feeding entirely on Heteroptera requires 4,500 fewer items than one eating only Coleoptera. For bobwhite chicks to attain normal growth rates, feeding trials conducted by Palmer (1995) suggest that a 7–10 day old chick requires approximately 6 g of invertebrates daily. In addition to protein, invertebrates also provide chicks with essential amino acids. The amino acids methionine and lysine have been identified as particularly important in feather development (Potts, 1986). Consequently, chicks that are unable to eat sufficient quantities of invertebrates suffer from poorer feather development as well as reduced growth rates (Southwood & Cross, 2002). The importance of providing invertebrate–rich brood rearing habitats for bobwhite has been recognised by land managers of game–shooting estates in southern Georgia and northern Florida for many years (Stoddard, 1931). Through the use of management prescriptions such as prescribed burning and cultivation of fallow fields, land managers annually create and maintain a patchwork of invertebrate–rich brood–rearing habitats across their estates. The provision of these and other habitats required by bobwhite has prevented population declines similar to those seen on agricultural landscapes (Brennan et al., 2000). To counteract the loss of bobwhite habitat on farmland in Georgia, farmers in selected areas are now able to enrol in the Bobwhite Quail Initiative (BQI), an agri– environmental scheme where payments are made for creating and managing habitats specifically for bobwhite (Cook, 2004). The establishment of brood–rearing cover along field margins is a key component of this scheme. While the number of farms that provide brood–rearing habitat through this scheme has increased, no studies have investigated the diet of bobwhite chicks foraging on this farmland. Here, the invertebrate composition in the diet of bobwhite chicks on farms enrolled in the BQI scheme was examined and then compared to the diet of chicks on a wild bobwhite shooting estate. Material and methods During spring 2002 and 2003, adult bobwhites were

Butler et al.

captured and fitted with a 6–g mortality sensing radio– transmitter on two sites in Georgia, United States. The first study site was located across two farms in central Georgia. Predominant crop types on both farms were cotton, peanuts, soya beans and maize. Both farms were participants in the BQI agri–environmental scheme and brood–rearing habitats had been established along field margins. The second site was a shooting estate intensively managed for wild bobwhite and other game species located in southern Georgia. The landscape is dominated by pine trees with an understory of grasses, forbaceous plants and shrubs. Between March and May each year, approximately 40–50% of the land area is burned in controlled fires. These fires encourage the growth of weedy vegetation that can harbour high densities of invertebrates (Hurst, 1972). In addition, small fields located across the estate, are also cultivated annually to encourage weed growth and create additional foraging areas for bobwhite broods. During the breeding season, April until September, nocturnal roosting sites (hereafter roost sites) of radio– collared adults with chicks were located until the brood was 14 days old. All chick–faecal matter found at each roost site was placed in a labelled plastic container and then frozen. Faecal samples were collected from 22 broods on the shooting estate and 19 broods on farmland. Analysis of faecal samples was conducted according to Moreby (1988). To account for differential recovery of diagnostic fragments from different invertebrates within a faecal sample, the proportion of each prey type in faecal samples was calculated using the formula described by Green & Tyler (1989) and correction factors described by Butler (2007). For each radioed–brood, the corrected data were pooled before the proportions of each invertebrate group in the diet were calculated. Statistical comparisons of these data were carried out using compositional analysis (Aebischer et al., 1993). As proportional data must sum to 1, the proportions are not linearly independent. To overcome this unit–sum constraint the proportional data can be converted to log–ratios. The log–ratios are independent of the category used as the denominator. To allow log–ratios to be calculated, all zero values are replaced by a very small proportion (0.001) (Aebischer et al., 1993). The log–ratio differences were calculated and tested simultaneously using MANOVA to reveal differences in the invertebrate composition between sites. If a significant difference was found, a ranking matrix was produced to determine where the differences lay (Aebischer et al., 1993). The differences between samples for all possible pairs of log–ratios were examined using t–tests. All analyses were conducted using Systat 8.0 (SPSS Inc., 1998). Results Using a six–part compositional analysis, relative differences in the proportions of Araneae, Hemiptera, Orthoptera, Hymenoptera, Coleoptera, and Others (predominately Lepidopteran larvae) in the diet of chicks from the two study sites were examined. The composition of these invertebrate groups in the diet of chicks varied between the two study sites (Λ = 0.590, F5, 35 = 4.858, P = 0.002)

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Table 1. Relative differences in the invertebrate composition in the diet of northern bobwhite chicks on a shooting estate and on farmland in Georgia, USA, 2002–2003. Invertebrate groups with a high rank were more abundant in the diet of broods on the shooting estate than in that of farmland broods. Different letters in the 'Ranks differ' column indicate an invertebrate group that differs significantly from another at P = 0.05. Invertebrate groups with the same letter do not differ significantly. Tabla 1. Diferencias relativas en la composición de invertebrados en la dieta de las crías septentrionales de colín de Virginia en un coto de caza y en cultivos de Georgia, EUA, 2002–2003. Los grupos de invertebrados con un número de rango alto fueron más abundantes en la dieta de las crías del coto de caza que en las de las tierras de labrantío. Las distintas letras de la columna "Ranks differ" indican que un grupo de invertebrados difiere significativamente de otro con P = 0,05. Los grupos de invertebrados con la misma letra no difieren significativamente.

Mean invertebrate composition (%)

Rank

Invertebrate group

Ranks differ

Shooting estate

Farmland

Coleoptera

43.8

26.2

Hymenoptera

19.0

17.2

Araneae

7.2

2.8

Orthoptera

7.3

6.9

Hemiptera

20.1

33.5

Others

2.6

13.4

(table 1). While > 44% of the invertebrates eaten by chicks on the shooting estate were beetles, this prey group accounted for only a quarter of the invertebrates in the diet of chicks foraging on farmland. Conversely, chicks foraging on farmland had eaten 1.7 times more Hemiptera and five times more Others than chicks on the shooting estate. Despite these differences in composition, Coleoptera, Hemiptera and Hymenoptera collectively formed approximately 80% of the invertebrate diet of bobwhite chicks on both sites. Discussion While chicks on both sites ate similar invertebrate groups, the composition of the diets differed. These differences are probably a reflection of the availability of prey items in the habitats used by broods on each landscape. Consistent with previous dietary studies, greater numbers of Hemiptera were found in the diet of chicks foraging on farmland than in forested landscapes (Palmer, 1995). Lepidoptera larvae were also five times more abundant in the diet of broods on the farmland. As shown in a companion radio–tracking study, the broods on the farmland site often used the brood–rearing habitats established under the BQI agri–environmental scheme, particularly the 6m non–sprayed headlands surrounding cropped fields (Cook, 2004). These types of habitats have been found to harbour high densities of these important chick–prey invertebrates (Rands, 1985; Chiverton & Sotherton, 1991; Palmer, 1995). The greater numbers of Coleoptera found in the diet of chicks on the shooting estate is consistent with the

results of other dietary studies where broods had been foraging in grass dominated habitats (Ford et al., 1938; Hurst, 1972). Large areas of wiregrass, Aristida stricta, were present on the shooting estate and many of the roost sites were located in or near these areas (Butler, 2007). The diet of broods on the shooting estate also consisted of 20% Hemiptera and 19% Hymenoptera, of which > 90% were Formicidae. While high numbers of Hemipetra have previously been reported in the diet of bobwhite chicks on shooting estates in this region, Formicidae have only previously been recorded as a trace item (Stoddard, 1931). The abundance of Formicidae in chick foraging habitats in this region may have increased over the last 70 years due to changes in habitat management techniques (Brennan, 1993) or because of the colonisation of the area by fire ants, Solenopsis spp. (Porter & Savignano, 1990). By increasing the availability of preferred prey items, the establishment of brood–rearing habitat on farmland has been shown to increase the proportion of these items in the diet of gamebird chicks (Sotherton et al., 1993). This, in turn, has also been found to increase chick–survival. The results of this study suggest that brood rearing habitats established under the BQI agri–environmental scheme provide bobwhite chicks with invertebrates known to be important dietary–items and could therefore improve chick survival rates on this landscape. However, due to the constraints of financial budgets and co–operation by farmers (Conover, 1998), it is difficult to envisage a sufficient quantity of BQI brood–rearing habitat being established on farmland in Georgia to reverse the dramatic declines of bobwhite populations seen over the last

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50 years (Brennan, 1991). Consequently, it is therefore important that the foraging–value of the cropped areas of arable fields is also improved. By also using crop management techniques such as conservation tillage with legume cover crops (Cederbaum et al., 2004) in conjunction with establishing brood–rearing habitats through BQI, farmers could vastly increase the availability of invertebrates in arable fields to bobwhite and other farmland birds. References Aebischer, N. J., Robertson, P. A. & Kenward, R. E., 1993. Compositional analysis of habitat use from animal radio–tracking data. Ecology, 74: 1313–1325. Butler, D. A., 2007. The role of invertebrates in the diet, growth and survival of northern bobwhite, Colinus virginianus, chicks in the southeastern United States. Ph. D. Thesis, Liverpool John Moores Univ. Brennan, L. A., 1991. How can we reverse the northern bobwhite population decline? Wildlife Society Bulletin, 19: 544–555. – 1993. Strip–disking: The forgotten bobwhite habitat management technique. Quail Unlimited Magazine, 12: 20–22. Brennan, L. A., Lee, J. M. & Fuller, R. S., 2000. Long– term trends of northern bobwhite populations and hunting success on private shooting plantations in northern Florida and southern Georgia. In: Quail IV: Proceedings of the Fourth National Quail Symposium: 75–77 (L. A. Brennan, W. E. Palmer, L. W. Burger, Jr & T. L. Pruden, Eds.). Tall Timbers Research Station, Tallahassee, Florida, USA. Cederbaum, S. B., Carroll, J. P. & Cooper, R. J., 2004. Effects of alternative cotton agriculture on avian and arthropod populations. Conservation Biology, 18: 1272–1282. Chiverton, P. A. & Sotherton, N. W., 1991. The effects on beneficial arthropods of the exclusion of herbicides from cereal crop edges. Journal of Applied Ecology, 28: 1027–1039. Conover, M. R., 1998. Perceptions of American agricultural producers about wildlife on their farms and ranches. Wildlife Society Bulletin, 26: 597–604. Cook, M. P., 2004. Northern bobwhite breeding

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season dispersal, habitat use, and survival in a southeastern agricultural landscape. M. S. Thesis, Univ. of Georgia. Ford, J., Chitty H. & Middleton, A. D., 1938. The food of Partridge Chicks (Perdix perdix) in Great Britain. Journal of Animal Ecology, 7: 251–265. Green, R. E. & Tyler, G. A., 1989. Determination of the diet of the stone curlew (Burhinus edicnemus) by faecal analysis. Journal of Zoology (London), 217: 311–320. Hurst, G. A., 1972. Insects and bobwhite quail brood habitat management. In: Proceedings First National Bobwhite Quail Symposium: 65–82 (J. A. Morrison & J. C. Lewis, Eds.). Oklahoma State Univ., Stillwater, Oklahoma, USA. Moreby, S. J., 1988. An aid to the identification of arthropod fragments in the faeces of gamebird chicks (Galliformes). Ibis, 130: 519–526. Palmer, W. E., 1995. Effects of modern pesticides and farming systems on northern bobwhite quail brood ecology. Ph. D. Thesis, North Carolina State Univ. Porter, S. D. & Savignano, D. A., 1990. Invasion of polygyne fire ants decimates native ants and disrupts arthropod community. Ecology, 71: 2095–2106. Potts, G. R., 1986. The Partridge: pesticides, predation and conservation. Collins, London, UK. Rands M. R. W., 1985. Pesticide use on cereals and the survival of grey partridge chicks: A field experiment. Journal of Applied Ecology, 22: 49–54. Stoddard, H. L., 1931. The Bobwhite Quail: its habits, preservation and increase. Charles Scribner’s Sons, New York. Stromborg, K. L., 1982. Modern pesticides and bobwhite populations. In: Proceedings Second National Bobwhite Quail Symposium: 69–73 (K. Schitoskey, Jr., E. C. Schitoskey & L. G. Talent, Eds.). Oklahoma State Univ., Stillwater, Oklahoma, USA. Sotherton, N. W., Robertson, P. A. & Dowell, S. D., 1993. Manipulating pesticide use to increase the production of wild game birds in Britain. In: Quail III: national quail symposium: 92–101 (K. E. Church & T. V. Dailey, Eds.). Kansas Dept of Wildlife and Parks, Pratt, Kansas, USA. Southwood, T. R. & Cross, D. J., 2002. Food requirements of grey partridge Perdix perdix chicks: Wildlife Biology, 8: 175–183.

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Microsatellite markers show distinctiveness of released and wild grey partridges in Finland T. Liukkonen, L. Kvist & S. Mykrä

Liukkonen, T., Kvist, L. & Mykrä, S., 2012. Microsatellite markers show distinctiveness of released and wild grey partridges in Finland. Animal Biodiversity and Conservation, 35.2: 419–428. Abstract Microsatellite markers show distinctiveness of released and wild grey partridges in Finland.— The main aim of this study was to study whether the present game farm stocks used for releases to the wild in Finland are similar to wild populations in their genetic structure, and if not, whether the wild populations show any signs of hybridisation. A total of 301 feather samples and ten microsatellite loci were used. Samples were collected from France, Great Britain, Finland (wild and captive) and Greece. We estimated pairwise FST–values between study populations, examined population structure and identified possible first generation migrants. Pairwise FST–values indicated structuring among studied populations. Results indicate that the farm stock used for releases deviates from the wild populations. No signs of hybridisation between the released and native birds were detected. Key words: Captive stock, Grey partridge, Microsatellites, Native stock, Perdix perdix. Resumen Los marcadores de microsatélites ponen de manifiesto las diferencias entre las perdices pardillas liberadas y silvestres en Finlandia.— El objetivo principal de este estudio consistió en estudiar si las poblaciones de las granjas cinegéticas utilizadas para las liberaciones en el medio natural en Finlandia son parecidas a las poblaciones silvestres en cuanto a su estructura genética y, en el caso de no serlo, si las poblaciones silvestres muestran signos de hibridación. Se utilizaron en total 301 muestras de pluma y 10 loci de microsatélite. Las muestras se recogieron en Francia, Gran Bretaña, Finlandia (silvestres y en cautividad) y Grecia. Calculamos los valores de FST entre pares de poblaciones del estudio, examinamos la estructura de la población y determinamos los posibles migrantes de primera generación. Los valores de FST entre pares indicaron la presencia de estructuración entre las poblaciones estudiadas. Los resultados indican que la población de granja utilizada para las liberaciones es distinta de las poblaciones silvestres. No se detectaron signos de hibridación entre las aves liberadas y las nativas. Palabras clave: Población en cautividad, Perdiz pardilla, Microsatélites, Población nativa, Perdix perdix. Received: 23 XII 11; Conditional acceptance: 19 II 12; Final acceptance: 10 IV 12 Tuija Liukkonen & Laura Kvist, Dept. of Biology, Univ. of Oulu, P. O. Box 3000, Fin–90014 Finland.– Sakari Mykrä, The Nature and Game Management Trust Finland, Makkarakoskentie 5A as1, 29600 Noormarkku, Finland.

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420

Liukkonen et al.

Introduction The distribution range of the grey partridge (Perdix perdix) covers large areas in Europe and Asia, all the way from Ireland to the Ural Mountains. The worldwide decline in the numbers of the grey partridge is well documented. A marked decline in the distribution range has occurred during the last century, mostly as a result of modern agricultural practices (for review, see Potts 1986). In Finland, the grey partridge lives at the edge of its northernmost distribution range. According to Kivirikko (1948), the grey partridge arrived in Finland from the southeast at the beginning of the 1800s, although the earliest observations were reported in 1690 (Merikallio, 1958). In 2007 the population size in Finland was estimated ca. 4,000 individuals and the species was classified as near–threatened (Liukkonen, 2007), but in the latest Finnish Red–List the species is classified as 'Least Concern' resulting from population size increase (Rassi et al., 2010). The first introductions were conducted in the middle of the 18th century (Merikallio, 1958) for hunting purposes with birds imported from Sweden (Kreuger, 1950). Captive–rearing and releasing of partridges has traditionally been carried out for game management purposes with the main aim to increase the size of the game bag. This kind of game management is and has been common for centuries. Early on, origin of stocks used for releases was rarely considered. At present, there are rules and recommendations for supplementing or replacing wild populations. In the IUCN Guidelines for the Re–introductions of Galliformes (WPA & IUCN/SSC Re–Introduction Specialist Group, 2009) it is stated that 'the sourcing of birds for re–introduction must not harm present populations and should be of appropriate (i.e. non– harmful) genetic stock. The taxonomic status of all remaining populations should be studied and, in most cases, the same subspecies or race should be used for reintroductions as those which were extirpated (unless adequate numbers are not available)'. The above–mentioned harmful effects are related to outbreeding depression, a phenomenon when matings between individuals from distinct populations break up co–adapted gene complexes and result in lower fitness of hybrid offspring. Examples of outbreeding depression range from plants and invertebrates to vertebrates and include reduction for example in viability, fertility, reproductive success and immune resistance (reviewed in Edmans, 2007). The European grey partridge is divided into two lineages by mitochondrial DNA (mtDNA). These lineages are assumed to refer to two different subspecies, P. p. perdix and P. p. lucida. After the last glaciation, colonisation of Europe occurred from two different glacial refugia, namely the Balkan Peninsula or Caucasus in the east and the Iberian Peninsula in the west (Liukkonen–Anttila et al., 2002). The western lineage, perdix, is widely found in Central Europe, for instance in France, Germany, Italy, Poland and the UK, whereas the eastern lineage, lucida, can be found in Finland, Greece, Bulgaria, Kazakhstan

Table 1. Collection locations, numbers and the assumed mtDNA–lineage of the grey partridge (Perdix perdix) feather samples used in this study. Tabla 1. Lugares y cifras de recolección y linaje esperado del ADNmt de las muestras de plumas de perdiz pardilla (Perdix perdix) utilizadas en este estudio.

Location

mtDNA–lineage

France Western 20 Great Britain Western 46 Finland, sites Western+ with releases eastern? 107 Finland, sites with no releases Eastern 54 Game farm stock Western+ eastern? 52 Eastern captive stock Eastern 7 Greece Eastern 15

and Ireland. It is possible, that at least in Estonia, Russia and Ukraine, populations are mixed, that is, either human–induced or naturally occurring, because birds of unknown origin have been released into these areas. In Finland, the native wild population represents the eastern mtDNA lineage, whereas most captive birds used for releases represent the western lineage (Liukkonen, 2006). This raises the question, have birds of wrong origin been released into the wild and have these releases had an impact on the wild population? Interest in managing grey partridge populations and willingness to conserve the native subspecies is not new in Finland. The Gene Bank Project has been going on for almost ten years. In this project the main aim has been to establish a general stock of eastern birds to be used in any possible releases to avoid mixing of these two lineages. The aims of this study were: 1) to study whether the game farm stock is similar to wild population in genetic structure and, 2) if not, to study whether the Finnish wild population shows signs of hybridisation between released and native birds. Materials and methods Sampled birds and laboratory methods Altogether, 301 feather samples of the grey partridge were used for this study (table 1). Samples were collected between 1998 and 2011 from Finland and, for comparisons, also from Great Britain, French Pyrenees and Greece (Liukkonen–Anttila et al., 2002; Liukkonen, 2006; this study; fig. 1). The Finnish wild

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Same farm stock Finland Eastern captive stock

Great Britain

France Greece

Fig. 1. Sampling locations of the grey partridges (Perdix perdix) used in this study. Fig. 1. Lugares de muestreo de las perdices pardilla (Perdix perdix) utilizadas en este estudio.

samples were collected from two types of sites; sites where no introductions have been made and sites where active releases for sport hunting take place. Samples from two captive populations were also obtained, one (game farm stock) is used for releases and the other (eastern captive stock) represents birds originating from the wild but not yet actively used for releases. DNA was extracted from feather quills as described in Liukkonen–Anttila et al. (2002) or by using QuickExtract solution (Epicentre) following the manufacturer’s protocol. Table 2 shows the microsatellite markers and modifications on PCR to amplify the loci used. The PCR products were run with ABI PRISM 3730 DNA Analyzer (Applied Biosystems) and scored using GeneMapper v. 4.0. Genetic variation Expected and observed heterozygosities were calculated with Arlequin v.3.11 (Excoffier et al., 2005). Allelic richness (corrected for the sample size bias with the rarefaction method) and inbreeding coefficients (FIS) were estimated with FSTAT v.2.9.3 (Goudet, 2001) excluding locus MNT408, because too few individuals successfully scored for this locus. Values were estimated for France, Great Britain (western subspecies), Finnish sites with no releases, Finnish sites with releases, game farm stock, eastern captive stock, and Greece (eastern subspecies).

Genetic structure Arlequin v.3.11 was used to estimate pairwise FST–values between the study populations. In addition, molecular variance analysis (AMOVA) was used to examine population structure using different groups defined a priori. Variation was estimated at three hierarchical levels; among groups (FCT), among populations within groups (FSC) and among populations (FST). Programme Structure v.2.2 (Pritchard et al., 2000; see also Falush et al., 2003) was used to infer the number of populations (K) in the data using the Markov chain Monte Carlo (MCMC) approach. A model with population admixture and correlated allele frequencies within populations (Falush et al., 2003) without prior information of the sampling locations was assumed. Five runs for each value of K between 1 and 12 were conducted, with a burn–in period of 100,000 iterations, and data were collected for 500,000 iterations. The likelihood of the data and following log probabilities for the different numbers of subpopulations were calculated for each K , the K with the highest log probability should equal the number of populations in the data. The largest change in log probability of data between consecutive numbers of populations, ΔK, has been proposed to estimate the actual K, and it has been found to perform better than the log probability per se (Evanno et al., 2005). This method should detect the highest level of population structure, when

422

Liukkonen et al.

Table 2. Microsatellite markers used for the analysis of the grey partridge (Perdix perdix) population structure. MgCL2 concentration (in mM) and annealing temperature (T, in ºC) are modifications of the original PCR protocols (Bech et al., 2010; Ferrero et al., 2007). Tabla 2. Marcadores de microsatélites utilizados para el análisis de la estructura de la población de perdiz pardilla (Perdix perdix). La concentración de MgCl2 (en mM) y la temperatura de hibridación (T, en ºC) son modificaciones de los protocolos originales de la PCR (Bech et al., 2010; Ferrero et al., 2007). MgCl2 T

Marker

Reference

Aru1A1

Ferrero et al., 2007

2.5 56

Aru1G4

Ferrero et al., 2007

2.5 56

Aru1E66

Ferrero et al., 2007

2.5 50

Aru1E102

Ferrero et al., 2007

2.0 50

Aru1F114

Ferrero et al., 2007

2.0 50

MNT12

Bech et al., 2010

2.0 53

MNT412

Bech et al., 2010

2.0 53

MNT477

Bech et al., 2010

2.0 55

MNT45

Bech et al., 2010

2.5 53

MNT408

Bech et al., 2010

2.5 53

several hierarchical levels exist, i.e. lower hierarchical structure may also be present. The results from Structure were used as input to this ad hoc method by Evanno et al. (2005). Factorial correspondence analysis (FCA) in the programme Genetix v. 4.0 (Belkhir et al., 2004) was used to visualise the relative similarity among samples and possible genetic structure within each region in a multivariate space. FCA tries to find the best linear combination of variables (i.e. allele frequencies at different loci in this case), which describe variation between individual observations. The factorial axes are ordered by their eigenvalues and the location of individuals is deﬁned according to the axis. The proximity of individuals along the axes expresses how genetically similar these individuals are. An assignment analysis and identification of possible first generation migrants between the sites was performed using the programme Geneclass 2 (Piry et al., 2004). This programme includes a Bayesian individual assignment method by Rannala & Mountain (1997) to estimate the marginal probability of each given individual genotype compared with the distribution of marginal probabilities of randomly generated genotypes (1,000 replicates) using the resampling method of Paetkau et al. (2004). We chose individuals that scored for at least four loci for this analysis. The

assignment threshold was set at 0.05 and alpha–level for the MCMC simulations was 0.01. Results The highest observed heterozygosities (table 3) were found in Greece and in the Finnish population with releases (0.748 and 0.710, respectively). The highest expected heterozygosities were found in the French population and again in the Finnish population with releases (0.759 and 0.763). These populations also harboured the highest allelic richness (2.905 and 2.811). The lowest observed heterozygosities were found in the British population and the Finnish game farm stock (0.537 and 0.610) and expected heterozygosities in the British and Greek populations (0.623 and 0.624). Inbreeding coefficients were significantly positive in the French population and in the Finnish sites with no releases (0.181 and 0.113; table 3). The pairwise FST values (table 4) between the study populations were almost all significant, with the exception that the Finnish sites with releases and sites with no releases did not differ from each other. The Greek population did not differ from the French population or from Finnish sites with no releases. The AMOVA analysis with different groupings yielded the highest FST–values when the British and French populations and Finnish game farm stock were grouped into one group and all the other Finnish populations with the Greek population (table 5). This grouping also resulted in the highest FCT–values (genetic difference among groups) and lowest FSC–values (difference among populations within groups). Results on the population structure suggested that the most likely number of populations would be five (mean Ln P(D) was –4373.7, for K = 4 mean Ln P(D) was –4382.86). By applying Evanno’s ΔK, the most likely number of populations was reduced to two (fig. 2A). The bar plots showing the proportion of each individual to belong to clusters indicate distinctiveness of the Finnish game farm stock from other Finnish samples. In addition, the bar plot for K = 2 suggests that this captive stock genetically belongs to the same cluster with most of the individuals from Great Britain and France (representing P. p. perdix). The individuals from the Finnish sites of releases and no releases and eastern captive stock were similar to the Greek individuals (representing P. p. lucida) (fig. 2B). The factorial correspondence analysis did not group populations into clearly distinct clusters. The individuals representing the eastern subspecies P. p. lucida and the Finnish sites and farm stocks were located 'in a pocket' within the individuals representing the western subspecies P. p. perdix indicating larger genetic variation in the western subspecies than in the other subspecies. The samples from Finnish sites with and without introductions and eastern farm stock tended to cluster together and separately from the Finnish game farm stock (fig. 3). The assignment test showed that no individual significantly deviated from the populations it was sampled from (table 5). However, in several cases

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Table 3. Observed (Ho) and expected (He) heterozygosities, allelic richness (A) and inbreeding coefficient (FIS) estimated from the studied populations. Standard deviations (SD) are given in the parentheses. Significant FIS–values (p < 0.05) are shown in bold. Tabla 3. Heterocigosis observada (Ho) y esperada (He), riqueza alélica (A) y coeficiente de endogamia (FIS) estimados de las poblaciones estudiadas. Las desviaciones estándar (SD) se indican entre paréntesis. Los valores significativos de FIS (p < 0,05) se muestran en negrita. Population

Ho (SD)

He (SD)

A (SD)

FIS

France

0.626 (0.262)

0.759 (0.170)

2.905 (0.585)

0.181

Great Britain

0.537 (0.251)

0.623 (0.212)

2.479 (0.663)

0.140

Finland, sites with no releases

0.685 (0.177)

0.730 (0.125)

2.697 (0.468)

0.113

Finland sites with releases

0.710 (0.178)

0.763 (0.103)

2.811 (0.416)

0.008

Game farm stock

0.610 (0.181)

0.643 (0.110)

2.395 (0.274)

0.080

Eastern captive stock

0.652 (0.145)

0.634 (0.176)

2.436 (0.573)

–0.029

Greece

0.748 (0.392)

0.624 (0.268)

2.430 (0.794)

–0.263

Table 4. Pairwise FST–values between the study populations: FR. Finnish sites with releases; FNR. Finnish sites with no releases; GFS. Game farm stock; ECS. Eastern captive stock. (Significant values with p < 0.05 are shown in bold.) Tabla 4. Valores de FST entre pares de poblaciones estudiadas: FR. Lugares finlandeses con liberaciones; FNR. Lugares finlandeses sin liberaciones; GFS. Población de granja cinegética; ECS. Población oriental en cautividad. (Los valores significativos con p < 0,05 se muestran en negrita.)

France

Great Britain

0.0525

0.0370

0.1162

FNR

0.0320

0.0708

–0.0168

GFS

0.0779

0.0933

0.1032

0.0853

ECS

0.1524

0.2339

0.0353

0.0705

0.1526

Greece

–0.0279

0.0420

0.0405

–0.0121

0.0536

the individuals yielded a higher assignment probability to belong to a population other than their own. These involved: Finnish sites with releases: one to Greece (N = 107); sites with no releases: nine to sites with releases (N = 54); game farm stock: seven to sites with releases, three to Great Britain, two to France and one to Finnish sites with no releases (N = 52); eastern captive stock: two to Finnish sites with releases, two to sites with no releases (N = 7). Only two possible first generation migrants were detected; one from Greece to the Finnish site with releases, and one from the site with releases to eastern captive stock. This observation merely reflects the affinities of these individuals to those populations and does not represent true migration events.

FNR

GFS

ECS

0.2056

Discussion Genetic variation The results of genetic variation obtained by using microsatellites were congruent with those obtained by mtDNA control region 1 sequences (Liukkonen–Anttila et al., 2002; Liukkonen, 2006). Expected heterozygosity and allelic richness were highest in the French population, whereas observed heterozygosity was highest in the Finnish sites with releases and in Greece. The lowest estimates of expected heterozygosity and allelic richness were obtained from Great Britain, Greece and the game farm stock, and the lowest observed heterozygosities were obtained from Great Britain and game farm stock.

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Table 5. AMOVA results using different groupings of the populations (GB. Great Britain; FR. Finnish sites with releases; FNR. Finnish sites with no releases; GFS. Game farm stock; ECS. Eastern captive stock): AG. Among groups; APG. Among populations within groups; WP. Within populations. (The grouping resulting to the highest FST– and FCT– and the lowest FSC–values is marked in italics and significant values with p < 0.001 are in bold.) Tabla 5. Resultados de AMOVA utilizando distintas agrupaciones de poblaciones (GB. Gran Bretaña; FR. Lugares finlandeses con liberaciones; FNR. Lugares finlandeses sin liberaciones; GFS. Población de granja cinegética; ECS. Población oriental en cautividad); AG. Entre grupos; APG. Entre poblaciones dentro de los grupos; WP. Dentro de las poblaciones. (La agrupación que tiene como resultado los valores mayores de FST y FCT y los menores para FSC está en cursiva y los valores significativos con p < 0,001 se muestran en negrita.) Grouping

APG

FST

FCT

FSC

GB+France/All Finnish+Greece

2.74

5.31

91.94

0.0806

0.0275

0.0546

GB+France+FR/FNR+GFS+ECS+Greece

–2.86

8.49

94.37

0.0563 –0.0286

0.0825

GB+France+FNR/FR+GFS+ECS+Greece

–1.59

7.64

93.94

0.0606 –0.0159

0.0752

GB+France+GFS/FR+FNR+ECS+Greece

4.67

3.66

91.68

0.0832

0.0467

0.0383

GB+France+ECS/FR+FNR+GFS+Greece

1.02

6.11

92.87

0.0713

0.0102

0.0617

GB+France+GFS+ECS/FR+FNR+Greece

3.18

4.57

92.25

0.0775

0.0318

0.0472

GB+France+GFS+ECS+FR/FNR+Greece

–3.34

8.26

95.08

0.0492 –0.0334

0.0799

GB+France+FR+FNR/GFS+ECS+Greece

1.96

5.63

92.41

0.0760

0.0196

0.0574

GB+France+FR+FNR+GFS+ECS/Greece

–4.44

7.42

97.01

0.0299 –0.0444

0.0711

The significantly positive inbreeding coefficients in the French and in the Finnish sites of no releases may result from small effective population sizes and isolation of these populations. French samples were collected from a small isolated area in the Pyrenees, indicating that the positive inbreeding coefficient may result from real inbreeding. However, the Finnish samples were collected from an area which is not geographically isolated. Thus, the positive inbreeding coefficient may also reflect the existence of an undetected population structure. The lowest observed heterozygosities were found in the British population and the Finnish game farm stock. The British population samples were from birds collected from the wild. The origin of these birds, however, is in translocated partridges. Translocation could have created a founder effect and loss of variation. The game farm stock in Finland has lost a great amount of genetic variation (Liukkonen, 2006), also likely resulting from an original founder effect followed by successive bottlenecks during breeding in captivity. Similar results on reduced amount of genetic variation were also found in the captive bred Mediterranean chukar partridge (Alectoris chukar, Barbanera et al., 2009a). Genetic structure Almost all pairwise FST–values between study populations were significant, thus supporting the previous results on the mtDNA control region 1 sequences

(Liukkonen–Anttila et al., 2002; Liukkonen, 2006). Samples from the wild in Finland grouped together with the Greek samples and the eastern captive birds, whereas Great Britain, France and the game farm stock grouped together. Finnish sites of releases and no releases did not differ from each other. As most introduced birds originate from the game farm stock, this indicates that released birds do not contribute to the wild population, but instead that their mortality after release might be high. The Greek population did not differ from the Finnish site of no releases, which supports the earlier results. In previous mtDNA studies, the Greek partridges represented the eastern lineage together with Finnish partridges (Liukkonen–Anttila et al., 2002; Liukkonen, 2006). However, against expectations, the Greek population did not differ from the French population either, possibly due to low sample sizes. Interestingly, the game farm stock, which is also used for releases, was clearly different from the wild population in Finland as well from the eastern captive stock. This result supported the earlier studies (Liukkonen–Anttila et al., 2002; Liukkonen, 2006). Also in the chukar partridge, farm stock in Crete differs from that used for releases (Barbanera et al., 2009b). The AMOVA analysis with different groupings yielded to the highest FST–values when the British and French populations and Finnish game farm stock were grouped into one group and all the rest of the Finnish populations with the Greek population. This grouping also resulted in the highest FCT–values (genetic difference

Animal Biodiversity and Conservation 35.2 (2012)

900

–4,100

800

–4,200

700

–4,300

ΔK

600

–4,400

500

–4,500

400

–4,600

300 200

–4,700

100

–4,800

0 B

K = 2

9 10

LnP (D)

425

–4,900

1 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0 K = 5

ca pt P. iv p. e lu ci da

fa rm

Ea st er n

G am e

no Sit re es le w as ith es

Si t re es le w as ith es

P. p. pe rd ix

1 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0

Fig. 2. Population structure of the grey partridge (Perdix perdix): A. The estimation suggested that the most likely number of populations would be five, but applying the Evanno's ΔK, the most likely number of populations was reduced to two; B. The bar plots on the grey partridge show the proportion of individuals belonging to different clusters. Fig. 2. Estructura de la población de perdiz pardilla (Perdix perdix): A. La estimación sugirió que el número más probable de poblaciones sería el de cinco, sin embargo al aplicar la ΔK de Evanno el número más probable de poblaciones se redujo a dos; B. los gráficos de barras de perdiz pardilla muestran la proporción de individuos pertenecientes a los diferentes grupos.

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Axe 2 (3.29%)

–1

–1 P. perdix perdix P. perdix lucida

0 Axe 1 (4.16%)

Sites with no releases Sites with releases Game farm stock Eastern captive stock

Fig. 3. The results of the factorial correspondence analysis on the grey partridge (Perdix perdix). The farm stock used for introductions shows distinctiveness from the wild populations. Fig. 3. Resultados del análisis factorial de correspondencia de perdiz pardilla (Perdix perdix). La población de granja utilizada para las introducciones muestra diferencias con las poblaciones silvestres.

among groups) and lowest FSC–values (difference among populations within groups). This, together with the results from structure analysis and factorial correspondence analysis, suggested that the game farm stock belongs to the same cluster with the western subspecies, whereas the native Finnish birds cluster with the eastern subspecies. If the game farm stock used for the releases into the wild is of the wrong origin, this is most likely detrimental to the natural population. It might lead to lowering fitness of the native population by breaking up adaptive gene complexes, especially if the released birds introduce traits that are adaptive in the environment they originate from but not in the environment they are released in. Releases of western subspecies of the grey partridge have been assumed to be one reason for the population crash, resulting from the differences in adaptation to cold environmental conditions in Finland (Siivonen, 1957). Captive–reared grey partridges and capercaillies are also known to clearly differ from their wild counterparts in several physiological and morphological traits (Putaala & Hissa, 1995; Pyörnilä et al., 1997; Liukkonen–Anttila et al., 2000), and this, too, may result in their low survival and contribution to the wild population. So far, however, no signs of hybridisation between the released and native birds could be detected with microsatellite markers. It is possible, that these

specific markers were not sensitive enough to reveal any hybrids. It is also possible that the native and the released partridges do not interbreed, that the released birds do not survive to breeding season (Puigcerver et al., 2007; Putaala et al., 2001), or that the possible hybrids have low fitness and disappear from the wild (Puigcerver et al., 2007). Released captive–bred red–legged and rock partridges (Alectoris graeca) are known to reproduce and hybridise in the wild (Randi, 2008). Introgressive hybridisation between wild local and captive released stocks might be threatening native populations by raising risks of outbreeding depression and loss of local adaptations. Massive translocations and releases of nonindigenous populations have threatened worldwide indigenous game bird populations as in the Italian grey partridge (P. p. italica, Liukkonen–Anttila et al., 2002), the common quail (Coturnix c. coturnix, Barilani et al., 2005) and the red–legged partridge (Tejedor et al., 2007; Randi, 2008; Barbanera et al., 2009b). Conclusions The Finnish native population seems to harbour quite a lot of genetic variation and it clusters together with individuals of the eastern subspecies, P. p. lucida. It is

Animal Biodiversity and Conservation 35.2 (2012)

evident that the game farm stock, which has been used for releases, deviates from this wild population. The birds from the eastern captive stock, which is derived from the wild, clustered together with the individuals from the native population and eastern subspecies. Microsatellite markers used in this study did not reveal any hybridisation between captive and wild populations. This finding supports the idea that 1) the game farm stock used for releases was of wrong origin and 2) luckily, no signs of hybridisation have yet been found between captive and native populations.. Whether this results from the lack of sensitivity among the used markers or is due to poor contribution of released grey partridges to the native populations remains to be solved in future studies. Acknowledgements We warmly thank all those who have sent us grey partridge feather samples during all these years, especially Dick Potts (UK), Claude Novoa (Pyrenees) and Bill Alexiou (Greece). Comments from two anonymous referees helped to improve the paper. This study was financially supported by the Nature and Game Management Trust Finland. References Barbanera, F., Guerrini, M., Khan, A. A. Panayides, P., Hadjigerou, P., Sokos, C., Gombobaatar, S., Samadi, S., Khan, B. Y., Tofanelli, S., Paoli, G. & Dini, F., 2009a. Human–mediated introgression of exotic chukar (Alectoris chukar, Galliformes) genes from East Asia into native Mediterranean partridges. Biological Invasions, 11: 333–348. Barbanera, F., Marchi, C., Guerrini, M., Panayides P., Sokos, C. & Hadjigerou, P., 2009b. Genetic structure of Mediterranean chukar (Alectoris chukar, Galliformes) populations: conservation and management implications. Naturwissenschaften, 96: 1203–1212. Barilani, M., Deregnaucourt, S. & Gallego S., Galli, L., Mucci, N., Piombo, R., Puigcerver, M., Rimondi, S., Rodríguez–Teijeiro, J. D., Spanò, S. & Randi, E., 2005. Detecting hybridization in wild (Coturnix c. coturnix) and domesticated (Coturnix c. japonica)quail populations. Biological Conservation, 126: 445–455. Bech, N., Novoa, C., Allienne, J. F. & Boissier, J., 2010. Transferability of microsatellite markers among economically and ecologically important galliform birds. Genetics and Molecular Research, 9: 1121–1129. Belkhir, K., Borsa, P., Chikhi, L., Raufaste, N. & Catch, F., 2004. GENETIX 4.0.5.2., Software under Windows™ for the genetics of the populations. Laboratory Genome, Populations, Interactions, CNRS UMR 5000, Univ. of Montpellier II, Montpellier, France. Edmans, S., 2007. Between a rock and a hard place: evaluating the relative risks of inbreeding and outbreeding for conservation and management.

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Molecular Ecology, 16: 463–475. Evanno, G., Regnaut, S. & Goudet, J., 2005. Detecting the number of clusters of individuals using the software Structure: A simulation study. Molecular Ecology 14: 2611–2620. Excoffier, L., Laval, G. & Schneider, S., 2005. Arlequin ver. 3.0: An integrated software package for population genetics data analysis. Evolutionary Bioinformatics Online, 1: 47–50. Falush, D., Stephens, M. & Pritchard, J. K., 2003: Inference of population structure using multilocus genotype data: linked loci and correlated allele frequencies. Genetics, 164: 1567–1587. Ferrero, M. E., González–Jara, P., Blanco–Aguiar, J. A., Sánchez–Barbudo, I. & Dávila, J. A., 2007. Sixteen new polymorphic microsatellite markers isolated for red–legged partridge (Alectoris rufa) and related species. Molecular Ecology Notes, 7: 1349–1351. Goudet, J., 2001. FSTAT, a program to estimate and test gene diversities and fixation indices (version 2.9.3). Available at http://www.unil.ch/izea/softwares/fstat.html. Kivirikko, K. E., 1948. Suomen linnut. WSOY, Porvoo. Finland. Kreuger, R., 1950. Om rapphönans, Perdix perdix, uppträdande in Finland under höstflyttningen. Ornis Fennica, 1–2: 48–51. Liukkonen, T., 2006. Finnish native grey partridge (Perdix perdix) population differs clearly in mitochondrial DNA from the farm stock used for releases. Annales Zoologici Fennici, 43: 271–279. – 2007. Peltopyykannan kehitys ja nykytila. In: Suomen peltopyykannan hoitosuunnitelma: 19–24 (MMM, Ed.) Maa–ja metsätalousministeriön julkaisuja 10/2007. [In Finnish.] Liukkonen–Anttila, T., Saartoala, R. & Hissa, R., 2000. Impact of hand–rearing on morphology and physiology of the capercaillie (Tetrao urogallus). Comparative Biochemistry and Physiology, 125A: 211–221. Liukkonen–Anttila, T., Uimaniemi, L., Orell, M. & Lumme, J., 2002. Mitochondrial DNA variation and phylogeography of the grey partridge (Perdix perdix) in Europe: from Pleistocene history to present day populations. Journal of Evolutionary Biology, 15: 971–982. Merikallio, E., 1958. Finnish Birds. Their distribution and numbers. Societas Pro Fauna et Flora Fennica. Fauna Fennica, V: 53–54. Paetkau, D., Slade, R., Burden, M. & Estoup, A., 2004. Genetic assignment methods for the direct, real–time estimation of migration rate: a simulation–based exploration of accuracy and power. Molecular Ecology, 13: 55–65. Piry, S., Alapetite, A., Cornuet, J–M., Paetkau, D., Baudouin, L. & Estoup, A., 2004. GENECLASS2: A software for genetic assignment and first–generation migrant detection. Journal of Heredity, 95: 536–539. Potts, G. R., 1986. The partridge: pesticides, predation and conservation. William Collins Sons & Co, London.

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Pritchard, J. K., Stephens, M. & Donnelly, P., 2000. Inference of population structure using multilocus genotype data. Genetics 155: 945–959. Puigcerver, M., Vinyoles, D. & Rodríguez–Teijeiro, J. D., 2007. Does restocking with Japanese quail or hybrids affect native populations of common quail Coturnix coturnix? Biological Conservation, 136: 628–635. Putaala, A. & Hissa, R., 1995. Effects of hand–rearing on physiology and anatomy in the grey partridge. Wildlife Biology, 1: 27–31. Putaala, A., Turtola, A. & Hissa R., 2001. Mortality of wild and released hand–reared grey partridges (Perdix perdix) in Finland. Game and wildlife science, 18: 291–304. Pyörnilä, A., Putaala, A. & Hissa, R., 1997. Fibre types in breast and leg muscles of hand–reared and wild grey partridge (Perdix perdix). Canadian Journal of Zoology, 76: 236–242. Randi, E., 2008. Detecting hybridization between wild species and their domesticated relatives. Molecular Ecology, 17: 285–293.

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Rannala, B. & Mountain, J. L., 1997. Detecting immigration by using multilocus genotypes. Proceedings of the National Academy of Sciences of the United States of America, 94: 9197–9201. Rassi, P., Hyvärinen, E., Juslén, A. & Mannerkoski, I. (Eds.), 2010. The 2010 Red List of Finnish Species. Ministry of the Environment and Finnish Environment Institute, Edita Ltd. Helsinki Siivonen, L., 1957. Peltopyy–ja rusakkokantojen vaihteluista ja niiden perussyistä sekä katojen torjumisesta. Suomen Riista, 11: 7–28. [In Finnish.] Tejedor, M. T., Monteagudo, L. V., Mautner, S., Hadjisterkotis, E. & Arruga, M. V., 2007.Introgression of Alectoris chukar genes into a Spanish wild Alectoris rufa population. Journal of Heredity, 98: 179–182. WPA (World Pheasant Association) & IUCN/SSC Re–introduction Specialist Group, (Eds.), 2009. Guidelines for the re–introduction of Galliformes for conservation purposes. IUCN. Gland, Switzerland. IUCN and Newcastle–upon–Tyne, UK: World Pheasant Association.

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Does the use of playback affect the estimated numbers of red–legged partridge Alectoris rufa? P. Tizzani, E. Negri, F. Silvano, G. Malacarne & P. G. Meneguz

Tizzani, P., Negri, E., Silvano, F., Malacarne, G., Meneguz, P. G., 2012. Does the use of playback affect the estimated numbers of red–legged partridge Alectoris rufa? Animal Biodiversity and Conservation, 35.2: 429–435. Abstract Does the use of playback affect the estimated numbers of red–legged partridge Alectoris rufa?— The red–legged partridge Alectoris rufa lives in a situation of potential conservation risk for its long–term preservation in Italy as its habitat is increasingly threatened by the disappearance of traditional agriculture–related environments. In such a situation, it is important to use effective and appropriate monitoring methods to assess population changes over time and to identify potential conservation threats. The objective of this study was to evaluate the effectiveness of the playback method to estimate the density of calling males. We compared playback method with spontaneous calling of males at dawn and direct observations along transects. The results on raw count data of playback counts revealed a strong underestimation rate compared to the method that gave the best results: count of spontaneous calls at dawn. Our study provides a critical evaluation of a method that is widely used even though data about its effectiveness are scarce. Our data do not evaluate detection probability of the three methods. Our aim was only to evaluate which methods give the best results in term of population size estimation under the same field condition (same population density, same period, same monitoring area). The results raise some doubts about the ability of the playback method to monitor red–legged partridge populations. The implications of our results for red–legged population management are discussed. Key words: Red–legged partridge, Census technique, Playback, Underestimation, Population monitoring, Raw count data. Resumen ¿Afecta el uso de playback a las cifras estimadas de perdiz roja, Alectoris rufa?— La perdiz roja, Alectoris rufa, vive en una situación de riesgo potencial en cuanto a su conservación a largo plazo en Italia, dado que su hábitat se ve cada vez más amenazado por la desaparición de los ambientes tradicionalmente relacionados con la agricultura. En tal situación, es importante aplicar métodos de control efectivos para estudiar los cambios poblacionales a través del tiempo, y la identificación de amenazas potenciales a la conservación. El objetivo de este estudio era evaluar la efectividad del método de playback para estimar la densidad de machos que vocalizaban. Comparamos el método del playback con las llamadas espontáneas de los machos al amanecer, y las observaciones directas a lo largo de transectos. El resultado de los datos sin procesar de los recuentos mediante playback revelaron una gran tasa de subestimación, comparados con el método que rindió los mejores resultados: el recuento de vocalizaciones espontáneas al amanecer. Nuestro estudio proporciona una evaluación crítica de un método que está ampliamente extendido, aunque los datos sobre su eficacia son escasos. Nuestros datos no evalúan la probabilidad de detección de los tres métodos. Nuestra intención era únicamente evaluar qué métodos arrojan los mejores resultados en términos de la estimación del tamaño de la población en las mismas condiciones de campo (misma densidad de población, mismo periodo, misma área de monitorización). Los resultados arrojan algunas dudas sobre la capacidad del método del playback para monitorizar las poblaciones de perdiz roja. Se discuten las implicaciones de nuestros resultados para su aplicación en la gestión de poblaciones de perdiz roja. Palabras clave: Perdiz roja, Método de censo, Playback, Subestimación, Monitorización poblacional, Datos de recuento sin procesar. ISSN: 1578–665X

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Received: 2 II 12; Conditional acceptance: 8 VI 12; Final acceptance: 19 IX 12 P. Tizzani & P. G. Meneguz, Dept. of Animal Production Epidemiology and Ecology, Univ. of Turin, Via Verdi 8, 10124 Torino, Italy.– E. Negri & F. Silvano, Stazzano Civic Museum of Natural History, Piazza Risorgimento 6, Stazzano, Italy.– G. Malacarne, Dept. of Environmental and Life Science, Univ. of Eastern Piedmont, 6 Via Duomo, Vercelli, Italy. Corresponding author: Paolo Tizzani. E–mail: paolo.tizzani@unito.it

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Introduction

Materials and methods

The red–legged partridge Alectoris rufa is widely distributed in Europe, with natural populations reported in Portugal, Spain, Andorra, France and Italy (Birdlife International, 2009). The IUCN Red List classifies this species as 'Least Concern', i.e., species without conservation threats (Birdlife International 2009). Red–legged partridge belong to the Galliform order, however, an order that is highly threatened, with 27% of species threatened with extinction, and about 60% of unthreatened species in decline (Rands, 1992; Potts & Aebischer, 1995; UNEP–WCMC, 2001). In Europe, the red–legged partridge is classified as SPEC 2 (Species of European Conservation Concern) for the following reasons: i) marked population decline in recent years and ii) population present only across Europe. The species has declined throughout its whole range, and it can be now considered vulnerable (Aebischer & Potts, 1994; Tucker & Heath, 1994; Aebischer & Lucio, 1996; Borralho et al., 1999). In Italy, in particular, the species showed a dramatic decline; it lives in a situation of potential conservation risk (Meriggi et al., 2007) for its long–term preservation as its habitat is increasingly threatened by the disappearance of traditional agriculture–related environments (Falcucci et al., 2006). In such a situation, it is important to use effective monitoring tools to assess population changes over time and to identify potential conservation threats (Kasprzykowski & Goławski, 2009). Many techniques are available to acquire estimates or relative indices of bird population sizes (e.g. Granholm, 1983; Verner, 1985; Zuberogoitia & Campos, 1998; Tryjanowski et al., 2003). But one of the most common methods is the point count of birds within hearing distance (Blondel et al., 1981). The problem related to this method is the uncertainty of detection probability of calling males, which is an important source of data variability (Pollock et al., 2002). The work of Jakob et al. (2010) established that the point count method with the use of song playback (playback call count) highly increases the detectability of calling males. This method is widely used to monitor many secretive species (e.g. Conway et al., 1993; Zuberogoitia & Campos, 1998; Brambilla & Rubolini, 2004) and in particular for Galliformes (e.g. Evans et al., 2007; Serrani et al., 2005; Cattadori et al., 2006; Amici et al., 2009). Our work aimed to evaluate the sensitivity of a census method widely used to monitor red legged partridge (Jakob et al., 2010). Even if it is considered a good method, as shown by Jakob et al. (2010), we wanted to test its effectiveness in the particular condition of a low density population, as occurs n Northern Italy where populations are declining (Meriggi & Mazzoni della Stella, 2004). It is important to consider that competition between males can be reduced at low densities, and so their urge to call could also be reduced (Lampe & Espmark, 1987; Penteriani et al., 2002; Penteriani, 2003). Our hypothesis was that in this case the method could be less precise. Spontaneous call, not being an expression of competition, would therefore be less affected by this problem.

Figure 1 shows the range of red–legged partridge in Italy, with a particular focus on the Alessandria Province (Piedmont Region). As for other species at the border of their distribution range (Hanski & Gaggiotti, 2004), the presence of low density populations is a normal condition. Our research was carried out in three study areas, each of 1,000 hectares, with the following characteristics: 1) hunting activity not allowed; 2) presence of low density populations, 3) particular conservation importance for red–legged partridge. The first study area was the typical Italian habitat for red–legged partridge, the traditional agricultural landscape. This habitat is seriously threatened by land abandonment and by the extension of wooded areas (Falcucci et al., 2006). In particular, this area is very important for conservation management of partridges because it is the only area in Italy where red–legged partridge populations are non–hybridized (Negri et al., in prep.). This area is located in the western part of Alessandria Province. The importance of the second and third study areas lies in the fact that in the last 10 years, red–legged partridge have colonized a completely new habitat, that it, the river bank, where it has never been reported before in Italy (Tizzani et al., 2011). So while the species is disappearing from its traditional habitat, it is progressively migrating to the North. These areas are located along the Scrivia River, one of the main rivers in the Alessandria province. Table 1 shows characteristic land use in the study areas. In view of the described situation, we have to survey populations at low density, either due to traditional habitat reduction, or to expansion in new, unusual habitat for Italy. We tested the sensitivity of the playback method along transects, as described by Jakob et al. (2010) and we compared the results to those obtained using two census methods: mapping calls at dawn (Pepin & Fouquet, 1992) and direct sighting along a transect (Borralho et al., 1996). Below we report the protocol used for the monitoring activities. Playback method The protocol is the same as that described in Jakob et al. (2010): (i) the playback is used along a transect; (ii) each transect consists of 8 playback stations; (iii) the playback stations are at least 500 m apart to avoid double counting; (iv) the playback repetition lasts from one hour before, to one hour after dawn; (v) the operators perform four calling sessions at each station for a total of 10 minutes; each session last two minutes and 40 seconds with a brief interval after each call; and (vi) this protocol is repeated on three consecutive days of the survey in April (peak of calling activity of males) in good meteorological conditions (no wind or rain). As control methods, first we mapped the spontaneous calling of males at dawn (Pepin & Foquet, 1992), This method exploits the spontaneous calling activity of males before dawn and localizes them on a map.

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Switzerland

Italy France

Piemonte

S lo ve ni a

Austria

Croatia

Alessandria

Monaco

San Merino

Vatican

Fig. 1. Red–legged partridge Italian range (in black) and study area (in light gray). Fig. 1. Distribución geográfica en Italia de perdiz roja (en negro), y área de estudio (en gris claro).

For mapping we used the same points that we used for playback calling, but in this case we needed more operators as all points had to be monitored at the same time. The spontaneous calling activity was recorded from 70 minutes before dawn till sunrise. The second control method was direct observation of partridges (from a car) along transects. Using this method all pairs seen on a transect in the same area of playback were recorded. After rain, more direct observations can be made because partridges come out to dry and they group on the roads. The transect method was applied along four transects (20 km) in area A, four transects (17 km) in area B, and three transects (13.4 km) in area C. The transects were monitored after rainfalls. Results and discussion Monitoring effort We monitored the three study areas for two years (2010 and 2011). Eleven transects were monitored each year for a total of 87 stations sampled three times (261 stations sampled). We monitored 31 stations in area A (sample area 608.4 ha), 32 stations in area B (628 ha) and 24 stations in area C (471 ha). The same stations were used for monitoring spontaneous calling. For direct observation we monitored transects for a total of 50.4 km. Playback results Data from playback monitoring showed very low male density in each study area, ranging between 1.48 males/100 ha and 3.78 males/100 ha (table 2).

Comparison of methods Comparing the number of partridges estimated using each of the three methods, spontaneous calling always showed the best results, followed by direct observation, then playback. We have to take into account that in our analysis we worked only with raw data, without correction for estimating detectability. In every case, considering that all data were obtained under similar field conditions, we can consider that the absolute comparison between methods is realistic. In particular, in figure 2 we report the method which gave the best results in each of the study areas in 2010 and 2011. The number of animals estimated with playback compared to numbers obtained with spontaneous calling varied between a minimum detection value of 34.6% of estimated animals to a maximum detection value of almost 70%. The underestimation value of playback decreases with the number of repetitions (fig. 3). From the first repetition to the last, there is, in fact, a strong increase in the number of new animals detected. This increase occurs above all between the first and the second repetition. This situation leads to i) a higher underestimation value if we consider the single repetition, and on the other hand, to ii) a low repeatability of the results (the coefficient of variation of the results is always very high, ranging from a minimum of 32% to a maximum of more than 100%, and it is always above the value considered reliable using standard census methods (Marchandeau et al., 2006). In our study the most important factors that affected repeatability of the playback method were the influence of daytime, and the attitude of males to reply to playback.

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Table 1. Land use characteristics of the study areas.

Table 2. Number of males and density (males/100 ha) of red–legged partridge in the three study areas.

Tabla 1. Características del uso del terreno de las áreas de estudio. Land use cover

Area A

Banks

Area B & C

11.6%

Cultivation

46.1%

60.3%

Pasture

1.9%

5.5%

Urban area

1.4%

1.7%

Vineyard

10.8%

Water Wooded area

Tabla 2. Número de machos y densidad (machos/100 ha) de perdiz roja en las tres áreas de estudio.

3.1%

39.8%

17.7%

Study area

Daytime can influence detectability because spontaneous calling activity varies throughout the day and the year, as demonstrated by Pepin & Fouquet (1992). In our study the peak activity was reached between 60 and 50 minutes before dawn (84% of spontaneous calling registered within the range). This trend iwas even reflected in the different number of induced calls (65 vs. 47) and their frequencies (135 vs. 85) recorded between the first and second hour of monitoring. From these observations it can be assumed that males have a higher probability of being detected in the first hour.

Calling 2011

9 (1.48)

23 (3.78)

10 (1.59)

11 (1.75)

10 (2.12)

12 (2.55)

Total

29 (1.70)

46 (2.69)

Daytime not only influences the birds’ response, but also the attitude of males to reply at playback. During our study only 10% of the males replied constantly to playback on all three monitoring days while a large proportion (almost 60%) replied only one time. This observation confirms that males of this species have a low tendency to call compared to other bird species (Eraud et al., 2007; Pagano & Arnold, 2009). Conclusion The results of our study confirm that in the case of low–density populations, the use of playback affects

2010

Calling 2010

2011

Callings/pairs

30 25 20

Spontaneous call

Direct observation Playback

10 5

A Area

Fig. 2. Comparison of the three count methods in the three study area. Data for spontaneous call in 2011 for area C, and for 2010 in area B and C are lacking. Fig. 2. Comparación de los tres métodos de recuento en las tres áreas de estudio. Faltan los datos de las vocalizaciones espontáneas en 2011 para el área C, y los de 2010 para las áreas B y C.

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2010

11 11

10 8

6 4

2011

2 0 First

Second

Third A

First

Second

Third

Fig. 3. Variation in number of calling males betweenh the three repetitions in the different study area. The three study areas are represented with different colours. Fig. 3. Variación entre las tres repeticiones en las distintas áreas de estudio, en el número de machos que emitían llamadas. Las tres áreas de estudio se representan con distintos colores.

the census estimates of red–legged partridge. This method can affect their estimation because raw data from monitoring a large proportion of males could not be detected using the count only. In low density populations, therefore, the use of spontaneous calls seems to be the best method with the lower underestimation value. Another valuable aspect of this method is that it can be used not only in spring but all year round, as spontaneous calling activity is not limited to a single season (Pepin & Fouquet, 1992). This method can thus be used to monitor male densities during spring and reproductive success of populations in August. A negative aspect is that many operators are needed to apply the method in large areas. Direct observation, however, can be a useful tool to integrate information derived from other methods as the male can be seen as well as heard and so it is possible to discriminate between paired and unpaired males. Even if playback highly increases detectability of calling males from red legged partridge (Jakob et al., 2010) and other partridge species (Kasprzykowski & Golawski, 2009; Novoa, 1992; Schoppers, 1996), the method shows a loss of sensibility in low density populations due to a higher incidence of confounding factors such as: i) lower attitude of males to reply to playback call; and ii) variation of calling activity during the day. Therefore, as although the method is good when detectability of males is high (Blondel et al., 1981) its capacity worsens as detectability decreases. Its limitations should be kept in mind when it is used to monitor population trends for low density and threatened populations. Nevertheless, it remains a useful tool to monitor large areas rapidly and with few operators.

References Aebischer, N. J. & Lucio, A. J., 1996. Red–legged Partridge Alectoris rufa. In: The EBCC Atlas of European Breeding Birds. Their distribution and abundance: 208–209 (W. J. M. Hagemeijer & M. J. Blair, Eds.). European Bird Census Council, T. & A. D. Poyser, London. Aebischer, N. J. & Potts, G. R., 1994. Red–legged Partridge Alectoris rufa. In: Birds in Europe, their conservation status. Birdlife Conservation Series, 3: 214–215 (G. M. Tucker & M. F. Heath, Eds.). BirdLife International, Cambridge. Amici, A., Pelorosso, R., Serrani, F. & Boccia, L., 2009. A nesting site suitability model for rock partridge (Alectoris graeca) in the Apennine Mountains using logistic regression. Ital. J. Anim. Sci., 8(2): 751–753. BirdLife International, 2009. Alectoris rufa. In: IUCN 2011. IUCN Red List of Threatened Species. Version 2011.2. <www.iucnredlist.org>. Downloaded on 01 February 2012. Blondel, J., Ferry, C. & Frochot, B., 1981. Point counts with unlimited distance. Stud. Avian biol., 6: 414–420. Borralho, R., Carvalho, S., Rego, F. & Vaz Pinto, P., 1999. Habitat correlates of red–legged partridge Alectoris rufa breeding density on Mediterranean farmland. Revue d’Écologie (La Terre et la Vie), 54: 59–69. Brambilla, M. & Rubolini, D., 2004. Water Rail Rallus aquaticus breeding density and habitat preferences in Northern Italy. Ardea, 92: 11–18.

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Cattadori, I. M., Ranci–Ortigosa, G., Gatto, M. & Hudson, P. J., 2006. Is the rock partridge Alectoris graeca saxatilis threatened in the Dolomitic Alps? Animal Conservation, 6(1): 71–81. Conway, C. J., Eddleman, W. R., Anderson, S. H. & Hene Bury, L. R., 1993. Seasonal changes in Yuma clapper rail vocalization rate and habitat use. Journal of Wildlife Management, 57: 282–290. Eraud, C., Boutin, J.–M., Roux, D. & Faivre, B., 2007. Spatial dynamics of an invasive bird species assessed using robust design occupancy analysis: the case of the Eurasian collared dove (Streptopelia decaocto) in France. J. Biogeogr., 34: 1077–1086. Evans, S. A., Mougeot, F., Redpath, S. M. & Leckie, F., 2007. Alternative methods for estimating density in an upland game bird: the red grouse Lagopus lagopus scoticus. Wildlife Biology, 13: 130–139. Falcucci, A., Maiorano, L. & Boitani, L., 2006. Changes in land–use/land–cover patterns in Italy and their implications for biodiversity conservation. Landscape Ecology, 22(4): 617–631. Granholm, S. L., 1983. Bias in density estimates due to movement of birds. Condor, 85: 243–248. Hanski, I. A. & Gaggiotti, O. E., 2004. Ecology, Genetics and Evolution of Metapopulations. Elsevier Academic Press, New York. Jakob, C., Ponce–Boutin, F., Besnard, A. & Eraud, C., 2010. On the efficiency of using song playback during call count surveys of Red–legged partridges (Alectoris rufa). European Journal of Wildlife Research, 56(6): 907–913. Kasprzykowski, Z. & Goławski, A., 2009. Does the use of playback affect the estimates of numbers of grey partridge Perdix perdix? Wildlife Biology, 15: 123–128. Lampe, H. M. & Espmark, Y. O., 1987. Singing activity and song pattern of the Redwing Turdus iliacus during the breeding season. Ornis Scandinavica, 18: 179–185. Marchandeau, S., Aubineau, J., Berger, F., Gaudin, J.–C., Roobrouck, A., Corda, E. & Reitz, F., 2006. Abundance indices: reliability testing is crucial – a field case of wild rabbit Oryctolagus cuniculus. Wildl. Biol., 12: 19–27. Meriggi, A. & Mazzoni della Stella, R., 2004. Dynamics of a reintroduced population of red–legged partridges Alectoris rufa in central Italy. Wildl. Biol., 9: 1–9. Meriggi, A., Mazzoni della Stella, R., Brangi, A., Ferloni, M., Masseroni, E., Merli, E. & Pompilio, L., 2007. The reintroduction of grey and red‐legged partridges (Perdix perdix and Alectoris rufa) in central Italy: a metapopulation approach. Italian Journal of Zoology, 74.3: 215–237. Negri, A., Pellegrino, I., Mucci, N., Randi, E., Tizzani, P. & Malacarne, G. (in prep.). Mitochondrial DNA and microsatellite markers demonstrate the occurrence of pure and chukar–introgressed Red–legged partridge (Alectoris rufa) populations in NW Italy. Novoa, C., 1992. Validation of a spring density index for Pyrenean grey partridge, Perdix perdix

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hispaniensis obtained with playbacks of recorded calls. Gibier Faune sauvage, 9: 105–118. Pagano, A. M. & Arnold, T. W., 2009. Estimating Detection Probabilities of Waterfowl Broods From Ground–Based Surveys. J. Wildlife Manage., 73(5): 686–694. Penteriani, V., 2003. Breeding density affects the honesty of bird vocal displays as possible indicators of male/territory quality. Ibis, 145(3): 127–135. Penteriani, V., Gallardob, M. & Cazassusc, H., 2002. Conspecific density biases passive auditory surveys. Journal of Field Ornithology, 73(4): 387–391. Pépin, D. & Fouquet, M., 1992. Factors affecting the incidence of dawn calling in red–legged and grey partridges. Behav. process, 26: 167–176. Pollock, K. H., Nichols, J. D., Simons, T. R., Farnsworth, G. R,. Bailey, L. L. & Sauer, J. R., 2002. Large scale wildlife monitoring studies: statistical methods for design and analysis. Environmetrics, 13: 1–15. Potts, G. R. & Aebischer, N. J., 1995. Population dynamics of the Grey Partridge Perdix perdix 1793–1993: monitoring, modelling and management. Ibis, 137: 29–37. Rands, M. R. W., 1992. The conservation status and priorities for threatened partridges, francolins and quails of the world. Gibier Faune Sauvage, 9: 493–502. Schoppers, J., 1996. Cassetterecorder goed hulpmiddel bij inventarisatie Patrijs Perdix perdix in het broedseizoen. Limosa, 69: 180–181. [In Dutch with an English summary.] Serrani, F., Sabatini, A., Amici, A., Fabiani, L. & Calò, C. M., 2005. A modified method of playback census for rock partridge (Alectoris graeca) in central Apennine, Italy: preliminary results. In: Proceedings of 4th International Symposium on Wild Fauna: 151. Tatranská Lomnica, Slovakia. Tizzani, P., Negri E., Silvano, F., Pellegrino, I., Malacarne, G. & Meneguz, P. G., 2011. Expansion of Red legged Partridge’s habitat in the southeast of the Alessandria Province: the rivers as ecological corridors? First Conference on Mediterranean population of the genus Alectoris Alessandria, Italy, 14–15 November 2011. Tryjanowski, P., Hromada, M., Antczak, M., Grzybek, J., Kuz´niak, S. & Lorek, G., 2003. Which method is most suitable for censusing breeding populations of red–backed (Lanius collurio) and great grey (Lanius excubitor) shrikes? Ornis Hungarica, 12–13: 223–228. Tucker, G. M. & Heath, M. F., 1994. Birds in Europe, their conservation status. Birdlife Conservation Series, 3. BirdLife International, Cambridge. UNEP–WCMC, 2001. Animals of the World Database. http://www.unep–wcmc.org/species/02–037 Merigianimals/animal_redlist. Verner, J., 1985. Assessment of counting techniques. Current Ornithology, 2: 247–302. Zuberogoitia, I. & Campos, L. F., 1998. Censusing owls in large areas: A comparison between methods. Ardeola, 45: 47–53.

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Animal Biodiversity and Conservation 35.2 (2012)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (abans Miscel·lània Zoològica) és una revista inter disciplinària publicada, des de 1958, pel Museu de Ciències Naturals de Barcelona. Inclou articles d'investigació empírica i teòrica en totes les àrees de la zoologia (sistemàtica, taxonomia, morfologia, biogeografia, ecologia, etologia, fisiologia i genètica) procedents de totes les regions del món amb especial énfasis als estudis que d'una manera o altre tinguin relevància en la biología de la conservació. La revista no publica compilacions bibliogràfiques, catàlegs, llistes d'espècies o cites puntuals. Els estudis realit zats amb espècies rares o protegides poden no ser acceptats tret que els autors disposin dels permisos corresponents. Cada volum anual consta de dos fascicles. Animal Biodiversity and Conservation es troba registrada en la majoria de les bases de dades més importants i està disponible gratuitament a internet a www.abc.museucienciesjournals.cat, de manera que permet una difusió mundial dels seus articles. Tots els manuscrits són revisats per l'editor execu tiu, un editor i dos revisors independents, triats d'una llista internacional, a fi de garantir–ne la qualitat. El procés de revisió és ràpid i constructiu. La publicació dels treballs acceptats es fa normalment dintre dels 12 mesos posteriors a la recepció. Una vegada hagin estat acceptats passaran a ser propietat de la revista. Aquesta es reserva els drets d’autor, i cap part dels treballs no podrà ser reproduïda sense citar–ne la procedència.

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El primer autor rebrà 50 separates del treball sense càrrec a més d'una separata electrònica en format PDF. Manuscrits Els treballs seran presentats en format DIN A–4 (30 línies de 70 espais cada una) a doble espai i amb totes les pàgines numerades. Els manuscrits han de ser complets, amb taules i figures. No s'han d'enviar les figures originals fins que l'article no hagi estat acceptat. El text es podrà redactar en anglès, castellà o català. Se suggereix als autors que enviïn els seus treballs en anglès. La revista els ofereix, sense cap càrrec, un servei de correcció per part d'una persona especialitzada en revistes científiques. En tots els casos, els textos hauran de ser redactats correctament i amb un llenguatge clar i concís. La redacció del text serà impersonal, i s'evitarà sempre la primera persona. Els caràcters cursius s’empraran per als noms científics de gèneres i d’espècies i per als neologis mes intraduïbles; les cites textuals, independentment de la llengua, seran consignades en lletra rodona i entre cometes i els noms d’autor que segueixin un tàxon aniran en rodona. Quan se citi una espècie per primera vegada en el text, es ressenyarà, sempre que sigui possible, el seu nom comú. Els topònims s’escriuran o bé en la forma original o bé en la llengua en què estigui escrit el treball, seguint sempre el mateix criteri. Els nombres de l’u al nou, sempre que estiguin en el text, s’escriuran amb lletres, excepte quan precedeixin una unitat de mesura. Els nombres més grans s'escriuran amb xifres excepte quan comencin una frase. Les dates s’indicaran de la forma següent: 28 VI 99 (un únic dia); 28, 30 VI 99 (dies 28 i 30); 28–30 VI 99 (dies 28 a 30). S’evitaran sempre les notes a peu de pàgina. Format dels articles Títol. Serà concís, però suficientment indicador del contingut. Els títols amb designacions de sèries numèriques (I, II, III, etc.) seran acceptats previ acord amb l'editor. Nom de l’autor o els autors Abstract en anglès que no ultrapassi les 12 línies mecanografiades (860 espais) i que mostri l’essència del manuscrit (introducció, material, mètodes, resultats i discussió). S'evitaran les especulacions i les cites bibliogràfiques. Estarà encapçalat pel títol del treball en cursiva. Key words en anglès (sis com a màxim), que orientin sobre el contingut del treball en ordre d’importància. Resumen en castellà, traducció de l'Abstract. De la traducció se'n farà càrrec la revista per a aquells autors que no siguin castellanoparlants. Palabras clave en castellà. © 2012 Museu de Ciències Naturals de Barcelona

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studies to conservation biology: a review. Anim. Biodivers. Conserv. La relació de referències bibliogràfiques d’un tre ball serà establerta i s’ordenarà alfabèticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c,... als treballs del mateix any. En el text, s’indicaran en la forma usual: "... segons Wemmer (1998)...", "...ha estat definit per Robinson & Redford (1991)...", "...les prospeccions realitzades (Begon et al., 1999)...". Taules. Es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Les taules grans seran més estretes i llargues que amples i curtes ja que s'han d'encaixar en l'amplada de la caixa de la revista. Figures. Tota classe d’il·lustracions (gràfics, figures o fotografies) entraran amb el nom de figura i es numeraran 1, 2, 3, etc. i han de ser sempre ressen yades en el text. Es podran incloure fotografies si són imprescindibles. Si les fotografies són en color, el cost de la seva publicació anirà a càrrec dels au tors. La mida màxima de les figures és de 15,5 cm d'amplada per 24 cm d'alçada. S'evitaran les figures tridimensionals. Tant els mapes com els dibuixos han d'incloure l'escala. Els ombreigs preferibles són blanc, negre o trama. S'evitaran els punteigs ja que no es reprodueixen bé. Peus de figura i capçaleres de taula. Seran clars, concisos i bilingües en la llengua de l’article i en anglès. Els títols dels apartats generals de l’article (Intro ducción, Material y métodos, Resultados, Discusión, Conclusiones, Agradecimientos y Referencias) no aniran numerats. No es poden utilitzar més de tres nivells de títols. Els autors procuraran que els seus treballs originals no passin de 20 pàgines (incloent–hi figures i taules). Si a l'article es descriuen nous tàxons, caldrà que els tipus estiguin dipositats en una institució pública. Es recomana als autors la consulta de fascicles recents de la revista per tenir en compte les seves normes.

Animal Biodiversity and Conservation 35.2 (2012)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (antes Miscel·lània Zoològica) es una revista inter disciplinar, publicada desde 1958 por el Museo Ciencias Naturales de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxonomía, mor fología, biogeografía, ecología, etología, fisiología y genética) procedentes de todas las regiones del mundo, con especial énfasis en los estudios que de una manera u otra tengan relevancia en la biología de la conservación. La revista no publica compila ciones bibliográficas, catálogos, listas de especies sin más o citas puntuales. Los estudios realizados con especies raras o protegidas pueden no ser aceptados a no ser que los autores dispongan de los permisos correspondientes. Cada volumen anual consta de dos fascículos. Animal Biodiversity and Conservation está re gistrada en todas las bases de datos importantes y además está disponible gratuitamente en internet en www.abc.museucienciesjournals.cat, lo que permite una difusión mundial de sus artículos. Todos los manuscritos son revisados por el editor ejecutivo, un editor y dos revisores independientes, elegidos de una lista internacional, a fin de garan tizar su calidad. El proceso de revisión es rápido y constructivo, y se realiza vía correo electrónico siem pre que es posible. La publicación de los trabajos aceptados se realiza con la mayor rapidez posible, normalmente dentro de los 12 meses siguientes a la recepción del trabajo. Una vez aceptado, el trabajo pasará a ser propie dad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reprodu cida sin citar su procedencia.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (abc@bcn.cat). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras y las tablas. Las figuras deberán enviarse también en archivos separados en formato TIFF, EPS o JPEG. Si se opta por la versión impresa, deberán remitirse cuatro copias juntamente con una copia en disquete a la Secretaría de Redacción. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre investigaciones originales no publicadas anteriormente y que se somete en ex clusiva a Animal Biodiversity and Conservation. En dicha carta también debe constar, para trabajos donde sea necesaria la manipulación de animales, que los autores disponen de los permisos necesa rios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores. Cuando el trabajo sea aceptado los autores de berán enviar a la Redacción una copia impresa de la versión final junto con un disquete del manuscrito ISSN: 1578–665X

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preparado con un procesador de textos e indican do el programa utilizado (preferiblemente Word). Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modificaciones sustanciales en las pruebas de im prenta, introducidas por los autores, irán a cargo de los mismos. El primer autor recibirá 50 separatas del trabajo sin cargo alguno y una copia electrónica en for mato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 líneas de 70 espacios cada una) a doble espacio y con las páginas numeradas. Los manuscritos de ben estar completos, con tablas y figuras. No enviar las figuras originales hasta que el artículo haya sido aceptado. El texto podrá redactarse en inglés, castellano o catalán. Se sugiere a los autores que envíen sus trabajos en inglés. La revista ofrece, sin cargo ningu no, un servicio de corrección por parte de una persona especializada en revistas científicas. En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso. La redacción del texto deberá ser impersonal, evitándose siempre la primera persona. Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común. Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo. Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase. Las fechas se indicarán de la siguiente forma: 28 VI 99 (un único día); 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30). Se evitarán siempre las notas a pie de página. Formato de los artículos Título. Será conciso pero suficientemente explicativo del contenido del trabajo. Los títulos con designacio nes de series numéricas (I, II, III, etc.) serán aceptados excepcionalmente previo consentimiento del editor. Nombre del autor o autores Abstract en inglés de 12 líneas mecanografiadas (860 espacios como máximo) y que exprese la esen cia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especula © 2012 Museu de Ciències Naturals de Barcelona

ciones y las citas bibliográficas. Irá encabezado por el título del trabajo en cursiva. Key words en inglés (un máximo de seis) que especifiquen el contenido del trabajo por orden de importancia. Resumen en castellano, traducción del abstract. Su traducción puede ser solicitada a la revista en el caso de autores que no sean castellano hablantes. Palabras clave en castellano. Dirección postal del autor o autores. (Título, Nombre, Abstract, Key words, Resumen, Palabras clave y Dirección postal conformarán la primera página.) Introducción. En ella se dará una idea de los ante cedentes del tema tratado, así como de los objetivos del trabajo. Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, me todología de estudio y análisis de los datos y zona de estudio. Resultados. En esta sección se presentarán úni camente los datos obtenidos que no hayan sido publicados previamente. Discusión. Se discutirán los resultados y se compara rán con otros trabajos relacionados. Las sugerencias sobre investigaciones futuras se podrán incluir al final de este apartado. Agradecimientos (optativo). Referencias. Cada trabajo irá acompañado de una bibliografía que incluirá únicamente las publicaciones citadas en el texto. Las referencias deben presentarse según los modelos siguientes (método Harvard): * Artículos de revista: Conroy, M. J. & Noon, B. R., 1996. Mapping of spe cies richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Libros y otras publicaciones no periódicas: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Trabajos de contribución en libros: Macdonald, D. W. & Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conserva tion biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford.

* Tesis doctorales: Merilä, J., 1996. Genetic and quantitative trait vari ation in natural bird populations. Tesis doctoral, Uppsala University. * Los trabajos en prensa sólo se citarán si han sido aceptados para su publicación: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Anim. Biodivers. Conserv. Las referencias se ordenarán alfabéticamente por autores, cronológicamente para un mismo autor y con las letras a, b, c,... para los trabajos de un mismo autor y año. En el texto las referencias bibliográficas se indicarán en la forma usual: "... según Wemmer (1998)...", "...ha sido definido por Robinson & Redford (1991)...", "...las prospecciones realizadas (Begon et al., 1999)...". Tablas. Se numerarán 1, 2, 3, etc. y se reseñarán todas en el texto. Las tablas grandes deben ser más estrechas y largas que anchas y cortas ya que deben ajustarse a la caja de la revista. Figuras. Toda clase de ilustraciones (gráficas, figuras o fotografías) se considerarán figuras, se numerarán 1, 2, 3, etc. y se citarán todas en el texto. Pueden incluirse fotografías si son imprescindibles. Si las fotografías son en color, el coste de su publicación irá a cargo de los autores. El tamaño máximo de las figuras es de 15,5 cm de ancho y 24 cm de alto. Deben evitarse las figuras tridimensionales. Tanto los mapas como los dibujos deben incluir la escala. Los sombreados preferibles son blanco, negro o trama. Deben evitarse los punteados ya que no se reproducen bien. Pies de figura y cabeceras de tabla. Serán claros, concisos y bilingües en castellano e inglés. Los títulos de los apartados generales del artículo (Introducción, Material y métodos, Resultados, Dis cusión, Agradecimientos y Referencias) no se nume rarán. No utilizar más de tres niveles de títulos. Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas. Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública. Se recomienda a los autores la consulta de fascículos recientes de la revista para seguir sus directrices.

Animal Biodiversity and Conservation 35.2 (2012)

Animal Biodiversity and Conservation

Manuscripts

Animal Biodiversity and Conservation (formerly Miscel·lània Zoològica) is an interdisciplinary journal published by the Natural Science Museum of Barce lona since 1958. It includes empirical and theoretical research from around the world that examines any aspect of Zoology (Systematics, Taxonomy, Morphol ogy, Biogeography, Ecology, Ethology, Physiology and Genetics). It gives special emphasis to studies related to Conservation Biology. The journal does not publish bibliographic compilations, listings, catalogues or collections of species, or isolated descriptions of a single specimen. Studies concerning rare or protected species will not be accepted unless the authors have been granted the relevant permits or authorisation. Each annual volume consists of two issues. Animal Biodiversity and Conservation is regis tered in all principal data bases and is freely available online at www.abc.museucienciesjournals.cat assuring world–wide access to articles published therein. All manuscripts are screened by the Executive Edi tor, an Editor and two independent reviewers so as to guarantee the quality of the papers. The review process aims to be rapid and constructive. Once accepted, papers are published as soon as is practicable. This is usually within 12 months of initial submission. Upon acceptance, manuscripts become the prop erty of the journal, which reserves copyright, and no published material may be reproduced or cited without acknowledging the source of information.

Manuscripts must be presented in DIN A–4 format, 30 lines, 70 keystrokes per page. Maintain double spacing throughout. Number all pages. Manuscripts should be complete with figures and tables. Do not send original figures until the paper has been accepted. The text may be written in English, Spanish or Cata lan, though English is preferred. The journal provides linguistic revision by an author’s editor. Care must be taken to use correct wording and the text should be written concisely and clearly. Scientific names of gen era and species as well as untranslatable neologisms must be in italics. Quotations in whatever language used must be typed in ordinary print between quota tion marks. The name of the author following a taxon should also be written in lower case letters. When referring to a species for the first time in the text, both common and scientific names should be given when possible. Do not capitalize common names of species unless they are proper nouns (e.g. Iberian rock lizard). Place names may appear either in their original form or in the language of the manuscript, but care should be taken to use the same criteria throughout the text. Numbers one to nine should be written in full within the text except when preceding a measure. Higher numbers should be written in numerals except at the beginning of a sentence. Specify dates as follows: 28 VI 99 (for a single day); 28, 30 VI 99 (referring to two days, e.g. 28th and 30th), 28–30 VI 99 (for more than two consecu tive days, e.g. 28th to 30th). Footnotes should not be used.

Information for authors Electronic submission of papers is encouraged (abc@bcn.cat). The preferred format is DOC or RTF. All figures must be readable by Word, embedded at the end of the manuscript and submitted together in a separate attachment in a TIFF, EPS or JPEG file. Tables should be placed at the end of the document. If a printed version is sent, four copies should be forwarded to the Editorial Office, together with a copy on computer disc. A cover letter stating that the article reports original research that has not been published elsewhere and has been submitted exclusively for consideration in Animal Biodiversity and Conservation is also necessary. When animal manipulation has been necessary, the cover letter should also specify that the authors follow current norms on the protec tion of animal species and that they have obtained all relevant permits and authorisations. Authors may suggest referees for their papers. Once an article has been accepted, authors should send a paper copy and an electronic copy of the final version. Please identify software (preferably Word). Proofs sent to the authors for correction should be returned to the Editorial Board within 10 days. Expenses due to any substantial alterations of the proofs will be charged to the authors. The first author will receive 50 reprints free of charge and an electronic version of the article in PDF format. ISSN: 1578–665X

Formatting of articles Title. Must be concise but as informative as possible. Numbering of parts (I, II, III, etc.) should be avoided and will be subject to the Editor’s consent. Name of author or authors Abstract in English, no longer than 12 typewritten lines (840 spaces), covering the contents of the article (introduction, material, methods, results and discussion). Speculation and literature citation should be avoided. The abstract should begin with the title in italics. Key words in English (no more than six) should express the precise contents of the manuscript in order of relevance. Resumen in Spanish, translation of the Abstract. Summaries of articles by non–Spanish speaking au thors will be translated by the journal on request. Palabras clave in Spanish. Address of the author or authors. (Title, Name, Abstract, Key words, Resumen, Palabras clave and Address should constitute the first page.) Introduction. Should include the historical back ground of the subject as well as the aims of the paper. © 2012 Museu de Ciències Naturals de Barcelona

Material and methods. This section should provide relevant information on the species studied, materi als, methods for collecting and analysing data, and the study area. Results. Report only previously unpublished results from the present study. Discussion. The results and their comparison with re lated studies should be discussed. Suggestions for future research may be given at the end of this section. Acknowledgements (optional). References. All manuscripts must include a bibliog raphy of the publications cited in the text. References should be presented as in the following examples (Harvard method): * Journal articles: Conroy, M. J. & Noon, B. R., 1996. Mapping of spe cies richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Books or other non–periodical publications: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Contributions or chapters of books: Macdonald, D. W. & Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conserva tion biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Ph. D. Thesis: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. Ph. D. Thesis, Uppsala University. * Works in press should only be cited if they have been accepted for publication: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Anim. Biodivers. Conserv. References must be set out in alphabetical and chrono

logical order for each author, adding the letters a, b, c,... to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to Wemmer (1998)...", "...has been defined by Robinson & Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." Tables. Must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal. Figures. All illustrations (graphs, drawings, photo graphs) should be termed as figures, and numbered consecutively in Arabic numerals (1, 2, 3, etc.) with reference in the text. Glossy print photographs, if essential, may be included. The Journal will publish colour photographs but the author will be charged for the cost. Figures have a maximum size of 15.5 cm wide by 24 cm long. Figures should not be tridimen sional. Any maps or drawings should include a scale. Shadings should be kept to a minimum and preferably with black, white or bold hatching. Stippling should be avoided as it may be lost in reproduction. Legends of tables and figures. Legends of tables and figures should be clear, concise, and written both in English and Spanish. Main headings (Introduction, Material and methods, Results, Discussion, Acknowledgements and Refer ences) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions.

Animal Biodiversity and Conservation 35.2 (2012)

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Payment method International cheque payable to Museu de Ciències Naturals de Barcelona and drawn against a Spanish bank Send cheque by postal mail to: Lluïsa Arroyo Dept. of Scientific Publications Nature Laboratory Museu de Ciències Naturals de Barcelona Psg. Picasso s/n. 08003 Barcelona, Spain Bank transfer to CaixaBank S. A. IBAN: ES 42 2100 3000 11 2201610475 SWIFT / BIC code: CAIXESBBXXX Send this order form by postal mail to:

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Animal Biodiversity and Conservation 35.2 (2012)

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Perdix 311–319 S. Faragó, G. Dittrich, K. Horváth–Hangya & D. Winkler Twenty years of the grey par tridge population in the LAJTA Project (Western Hungary) 321–331 P. J. K. McGowan, L. L. Owens & M. J. Grainger Galliformes science and species extinctions: what we kwow and what we need to know 333–342 J. D. Rodríguez–Teijeiro, F. Sardà–Palomera & M. Puigcerver Post–breeding movements and migration patterns of western populations of common quail (Coturnix coturnix): from knowledge to hunting management 343–352 M. Puigcerver, F. Sardà–Palomera & J. D. Rodríguez– Teijeiro Deter mining population tr ends and conservation status of the common quail (Coturnix coturnix) in Western Europe 353–362 N. J. Aebischer & J. A. Ewald The grey partridge in the UK: population status, research, policy and prospects

363–369 J. A. Ewald, G. R. Potts & N. J. Aebischer Restoration of a wild grey partridge shoot: a major development in the Sussex study, UK 371–380 V. A. Bontzorlos, C. G. Vlachos, D. E. Bakaloudis, E. N. Chatzinikos, E. A. Dedousopoulou, D. K. Kiousis & C. Thomaides Rock partridge (Alectoris graeca graeca) population density and trends in central Greece 381–386 R. A. H. Draycott Restoration of a sustainable wild grey partridge shoot in eastern England 387–393 K. Buckley, P. Kelly, B. Kavanagh, E. C. O’Gorman, T. Carnus & B. J. McMahon Every partridge counts, successful techniques used in the captive conservation breeding programme for wild grey partridge in Ireland 395–404 A. Mateo–Moriones, R. Villafuerte & P. Ferreras Does fox control improve red–legged partridge (Alectoris rufa) survival? An experimental study in Northern Spain 405–413 E. Bro, P. Mayot & F. Reitz Effectiveness of habitat management for improving grey partridge populations: a BACI experimental assessment 415–418 D. A. Butler, W. E. Palmer & M. P. Cook The invertebrate diet of northern bobwhite chicks in Georgia, United States 419–428 T. Liukkonen, L. Kvist & S. Mykrä Microsatellite markers show distinctiveness of released and wild grey partridges in Finland 429–435 P. Tizzani, E. Negri, F. Silvano, G. Malacarne & P. G. Meneguz Does the use of playback affect the estimated numbers of red–legged partridge Alectoris rufa

Les cites o els abstracts dels articles d’Animal Biodiversity and Conservation es resenyen a / Las citas o los abstracts de los artículos de Animal Biodiversity and Conservation se mencionan en / Animal Biodiversity and Conservation is cited or abstracted in: Abstracts of Entomology, Agrindex, Animal Behaviour Abstracts, Anthropos, Aquatic Sciences and Fisheries Abstracts, Behavioural Biology Abstracts, Biological Abstracts, Biological and Agricultural Abstracts, BIOSIS Previews, Current Primate References, Current Contents/Agriculture, Biology & Environmental Sciences, DIALNET, DOAJ, DULCINEA, e–revist@s, Ecological Abstracts, Ecology Abstracts, Entomology Abstracts, Environmental Abstracts, Environmental Periodical Bibliography, Genetic Abstracts, Geographical Abstracts, Índice Español de Ciencia y Tecnología, International Abstracts of Biological Sciences, International Bibliography of Periodical Literature, International Developmental Abstracts, Latindex, Marine Sciences Contents Tables, Oceanic Abstracts, RACO, Recent Ornithological Literature, Referatirnyi Zhurnal, Science Abstracts, Science Citation Index Expanded, Scientific Commons, SCImago, SCOPUS, Serials Directory, SHERPA/RoMEO, Ulrich’s International Periodical Directory, Zoological Records.

Índex / Índice / Contents Animal Biodiversity and Conservation 35.2 (2012) ISSN 1578–665X 153–154 F. Buner & M. Puigcerver XXXth IUGB Congress and Perdix XIII

209–217 C. Rosell, F. Navàs & S. Romero Reproduction of wild boar in a cropland and coastal wetland area: implications for management

IUGB

219–220 C. Rosell & F Llimona Human–wildlife interactions

159–161 S. Bertouille Wildlife law and policy 163–170 K. G. Papaspyropoulos, J. Koufis, L. Tourlida & A. Georgakopoulou Estimating the economic impact of a long term hunting ban on local businesses in a rural areas in Greece a hypothetical scenario

171–174 M. B. Ellis Management of waterfowl shooting during periods of severe weather in the UK

235–246 E. Belotti, M. Heurich, J. Kreisinger, P. Šustr & L. Bufka Influence of tourism and traffic on the Eurasian lynx hunting activity and daily movements

175–188 T. Beroud, J. Druais, Y. Bay & J. C. Ricci Visual counts, bioacoustics and RADAR: three methods to study waterfowl prenuptial migration in Southern France

247–252 V. J. Colino–Rabanal, J. Bosch, Mª J. Muñoz & S. J. Peris Influence of new irrigated croplands on wild boar (Sus scrofa) road kills in NW Spain

189–196 C. Fischer & R. Tagand S p at i a l b e h av i o u r a n d s u r v i v a l translocated wild brown hares

253–265 K. Morelle, P. Bouché, F. Lehaire, V. Leeman & P. Lejeune Game species monitoring using road–based distance sampling in association with thermal imagers: a covariate analysis

Consorci format per: Amb el suport de